Ecotoxicology of amphibians and reptiles [2nd ed] 9781420064179, 8504691500, 1420064177

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Ecotoxicology of Amphibians and Reptiles Second Edition

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Other Titles from the Society of Environmental Toxicology and Chemistry (SETAC) Other Titles from the Society of Environmental Toxicology and Chemistry (SETAC) Ecological Assessment of Selenium in the Aquatic Environment Adams, Brooks, Linking Aquatic Chapman, Exposure and Effects: Risk editors Assessment of Pesticides Brock, Alix, Brown, Capri, Gottesbüren,2010 Heimbach, Lythgo, Schulz, Streloke, editors 2009 Application of Uncertainty Analysis to Ecological Risk of Pesticides Derivation and Use of Environmental QualityHart, and Human Warren-Hicks, editors Health Standards for Chemical Substances 2010 in Water and Soil Crane, Matthiessen, Maycock, Merrington, Whitehouse, editors 2009 Risk Assessment of Pesticides Linking Aquatic Exposure and Effects: Brock, Alix, Brown, Capri, Gottesbüren, Heimbach, Lythgo, Schulz, Streloke, editors Aquatic Macrophyte Risk Assessment for Pesticides 2010 Maltby, Arnold, Arts, Davies, Heimbach, Pickl, Poulsen, editors 2009 Ecological Models for Regulatory Risk Assessments of Pesticides: Developing a Strategy for the Future Thorbek, Forbes, Heimbach, Hommen, Thulke, Van den Brink, Wogram, Grimm, editors Veterinary Medicines 2010in the Environment Crane, Boxall, Barrett Derivation and Use of Environmental Quality and 2008 Human Health Standards for Chemical Substances in Water and Soil RelevanceCrane, of Ambient Water Quality Criteria for Ephemeral and Effluent-dependent Matthiessen, Maycock, Merrington, Whitehouse, editors Watercourses of the Arid Western United States 2010 Gensemer, Meyerhof, Ramage, Curley 2008 Aquatic Macrophyte Risk Assessment for Pesticides Maltby, Arnold, Arts, Davies, Heimbach, Pickl, Poulsen Extrapolation Practice for Ecotoxicological Effect Characterization of Chemicals 2010 Solomon, Brock, de Zwart, Dyev, Posthumm, Richards, editors 2008 Veterinary Medicines in the Environment Crane, Boxall, Barrett Environmental Life Cycle Costing 2008 Hunkeler, Lichtenvort, Rebitzer, editors 2008 Relevance of Ambient Water Quality Criteria for Ephemeral and Effluent dependent Watercourses of the Arid Western United States Valuation of Ecological Resources: Integration of Ecology and Socioeconomics Gensemer, Meyerhof, Ramage,Making Curley in Environmental Decision Stahl, Kapustka, 2008 Munns, Bruins, editors 2007 For information about SETAC publications, including SETAC’s international journals, Environmental Toxicology and Chemistry and Integrated Environmental Assessment and Management, contact the SETAC office nearest you: SETAC 1010 North 12th Avenue Pensacola, FL 32501-3367 USA T 850 469 1500 F 850 469 9778 E [email protected]

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Ecotoxicology of Amphibians and Reptiles Second Edition Edited by

Donald W. Sparling, Greg Linder, Christine A. Bishop, Sherry K. Krest

Coordinating Editor of SETAC Books Joseph W. Gorsuch Copper Development Association, Inc. New York, NY, USA

Boca Raton London New York

CRC Press is an imprint of the Taylor & Francis Group, an informa business

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Information contained herein does not necessarily reflect the policy or views of the Society of Environmental Toxicology and Chemistry (SETAC). Mention of commercial or noncommercial products and services does not imply endorsement or affiliation by the author or SETAC.

Published in collaboration with the Society of Environmental Toxicology and Chemistry (SETAC) 1010 North 12th Avenue, Pensacola, Florida 32501 Telephone: (850) 469-1500 ; Fax: (850) 469-9778; Email: [email protected] Web site: www.setac.org © 2010 by the Society of Environmental Toxicology and Chemistry (SETAC) SETAC Press is an imprint of the Society of Environmental Toxicology and Chemistry. No claim to original U.S. Government works Printed in the United States of America on acid-free paper 10 9 8 7 6 5 4 3 2 1 International Standard Book Number-13: 978-1-4200-6417-9 (Ebook-PDF) This book contains information obtained from authentic and highly regarded sources. Reasonable efforts have been made to publish reliable data and information, but the author and publisher cannot assume responsibility for the validity of all materials or the consequences of their use. The authors and publishers have attempted to trace the copyright holders of all material reproduced in this publication and apologize to copyright holders if permission to publish in this form has not been obtained. If any copyright material has not been acknowledged please write and let us know so we may rectify in any future reprint. Except as permitted under U.S. Copyright Law, no part of this book may be reprinted, reproduced, transmitted, or utilized in any form by any electronic, mechanical, or other means, now known or hereafter invented, including photocopying, microfilming, and recording, or in any information storage or retrieval system, without written permission from the publishers. For permission to photocopy or use material electronically from this work, please access www.copyright.com (http:// www.copyright.com/) or contact the Copyright Clearance Center, Inc. (CCC), 222 Rosewood Drive, Danvers, MA 01923, 978-750-8400. CCC is a not-for-profit organization that provides licenses and registration for a variety of users. For organizations that have been granted a photocopy license by the CCC, a separate system of payment has been arranged. Trademark Notice: Product or corporate names may be trademarks or registered trademarks, and are used only for identification and explanation without intent to infringe. Visit the Taylor & Francis Web site at http://www.taylorandfrancis.com and the CRC Press Web site at http://www.crcpress.com and the SETAC Web site at www.setac.org

SETAC Publications Books published by the Society of Environmental Toxicology and Chemistry (SETAC) provide in-depth reviews and critical appraisals on scientific subjects relevant to understanding the impacts of chemicals and technology on the environment. The books explore topics reviewed and recommended by the Publications Advisory Council and approved by the SETAC North America, Latin America, or Asia/Pacific Board of Directors; the SETAC Europe Council; or the SETAC World Council for their importance, timeliness, and contribution to multidisciplinary approaches to solving environmental problems. The diversity and breadth of subjects covered in the series reflect the wide range of disciplines encompassed by environmental toxicology, environmental chemistry, hazard and risk assessment, and life-cycle assessment. SETAC books attempt to present the reader with authoritative coverage of the literature, as well as paradigms, methodologies, and controversies; research needs; and new developments specific to the featured topics. The books are generally peer reviewed for SETAC by acknowledged experts. SETAC publications, which include Technical Issue Papers (TIPs), workshop summaries, newsletter (SETAC Globe), and journals (Environmental Toxicology and Chemistry and Integrated Environmental Assessment and Management), are useful to environmental scientists in research, research management, chemical manufacturing and regulation, risk assessment, and education, as well as to students considering or preparing for careers in these areas. The publications provide information for keeping abreast of recent developments in familiar subject areas and for rapid introduction to principles and approaches in new subject areas. SETAC recognizes and thanks the past coordinating editors of SETAC books: A.S. Green, International Zinc Association Durham, North Carolina, USA C.G. Ingersoll, Columbia Environmental Research Center US Geological Survey, Columbia, Missouri, USA T.W. La Point, Institute of Applied Sciences University of North Texas, Denton, Texas, USA B.T. Walton, US Environmental Protection Agency Research Triangle Park, North Carolina, USA C.H. Ward, Department of Environmental Sciences and Engineering Rice University, Houston, Texas, USA

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Cover photo credits Sampling in the Sierras: With permission from Deborah F. Cowman, Executive Director, Brazos Valley Museum of Natural History, Bryan, Texas. Western fence lizard: With permission from Bill Bouton, http://www.flickr.com/photos/billbouton /sets Pacific tree frog skeletal view: With permission from Brandon Ballengée, “The Complex LifeCycle of the Trematode, Riberoria ondatrae” by Brandon Ballengée, 2002, 11 by 14 inches. Sakura Ink on water-colour paper. Courtesy the artist and Archibald Arts, New York, NY. Collection of Anthony Archibald J. Figure is from Rohr, J.R., T. Raffel, and S.K. Sessions. (2008). Parasites and Amphibians. Chapter 4. In: Amphibian Biology, Conservation and Decline of Amphibians (H, Heatwole, ed.) Chipping Norton, Australia: Surrey Beatty & Sons. Pacific tree frog: By Gary Fellers, US Geological Survey, copyright free. European pond turtle: With permission from Manuel Ortiz Santaliestra, Instituto de Investigación en Recursos Cinegéticos, Ciudad Real, Spain.

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Contents List of Figures....................................................................................................................................ix List of Tables................................................................................................................................... xiii About the Editors............................................................................................................................xvii List of Contributors..........................................................................................................................xix Acknowledgments............................................................................................................................xxi Preface for Second Edition........................................................................................................... xxiii Preface from the First Edition........................................................................................................xxv Chapter 1. Recent Advancements in Amphibian and Reptile Ecotoxicology................................1 Donald W. Sparling, Greg Linder, Christine A. Bishop, and Sherry K. Krest Chapter 2. Declines and the Global Status of Amphibians.......................................................... 13 Ross A. Alford Chapter 3. The Global Status of Reptiles and Causes of Their Decline....................................... 47 Brian D. Todd, John D. Willson, and J. Whitfield Gibbons Chapter 4. Ecotoxicology of Amphibians and Reptiles in a Nutshell.......................................... 69 Greg Linder, Christine M. Lehman and Joseph R. Bidwell Chapter 5. Physiological Ecology of Amphibians and Reptiles: Natural History and Life History Attributes Framing Chemical Exposure in the Field.................... 105 Greg Linder, Brent D. Palmer, Edward E. Little, Christopher L. Rowe, and Paula F.P. Henry Chapter 6. Effects of Current-Use Pesticides on Amphibians.................................................... 167 Christine M. Lehman and Bethany K. Williams Chapter 7. Ecotoxicology of Pesticides in Reptiles....................................................................203 Bruce D. Pauli, Stacey Money, and Donald W. Sparling Chapter 8. Atrazine in the Environment and Its Implications for Amphibians and Reptiles..... 225 Christine A. Bishop, Tana V. McDaniel, and Shane R. de Solla Chapter 9. Ecotoxicology of Organic Contaminants to Amphibians......................................... 261 Donald W. Sparling

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Contents

Chapter 10. Organic Contaminants in Reptiles............................................................................ 289 Shane R. de Solla Chapter 11. Interdisciplinary and Hierarchical Approaches for Studying the Effects of Metals and Metalloids on Amphibians................................................................. 325 W.A. Hopkins and Christopher. L. Rowe Chapter 12. The Ecotoxicology of Metals in Reptiles.................................................................. 337 Britta Grillitsch and Luis Schiesari Chapter 13. Solar UV Radiation and Amphibians: Factors Mitigating Injury.............................449 Edward E. Little and Robin D. Calfee Chapter 14. Multiple Stressors and Indirect Food Web Effects of Contaminants on Herptofauna.......................................................................................................... 475 Rick A. Relyea Chapter 15. Emerging Contaminants and Their Potential Effects on Amphibians and Reptiles............................................................................................................... 487 Laura L. McConnell and Donald W. Sparling Chapter 16. A Decade of Deformities: Advances in Our Understanding of Amphibian Malformations and Their Implications.............................................. 511 Pieter T.J. Johnson, Mari K. Reeves, Sherry K. Krest, and Alfred E. Pinkney Chapter 17. Population Estimation Methods for Amphibians and Reptiles................................. 537 Larissa L. Bailey and Marc J. Mazerolle Chapter 18. Epilogue: Ecotoxicology of Amphibians and Reptiles — Where Should We Be Going and How Do We Get There?............................................................... 547 Greg Linder, Christine A. Bishop, Sherry K. Krest, and Donald W. Sparling Appendix: Metal Contamination in Reptiles: An Appendix of Data Compiled from the Existing Literature....................................................................................................... 553 Britta Grillitsch and Luis Schiesari Index...............................................................................................................................................905

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List of Figures Figure 1.1

Total number of contaminant-related papers published between 1996 and 2008 by vertebrate class..................................................................................3

Figure 1.2

Annual numbers of scientific publications for amphibians and reptiles between 1996 and 2008 as listed in ISI Web of Science........................................4

Figure 1.3

Number of contaminant-related papers published for amphibians and reptiles between 1996 and 2008.............................................................................5

Figure 1.4

Percent of all scientific publications for amphibians and reptiles between 1996 and 2008 that were contaminant related. . ................................................ 5

Figure 1.5

Contaminant-related papers published on amphibians between 1996 and 2008 by chemical class ...................................................................................6

Figure 1.6

Contaminant-related papers for amphibians that were published between 1996 and 2008 by taxonomic group.......................................................................6

Figure 1.7

A comparison of the numbers of papers published for amphibians between 1996 and 2008 by associated stressor.....................................................................7

Figure 1.8

Number of contaminant-related papers published between 1996 and 2008 on reptiles by type of contaminant.........................................................................7

Figure 1.9

Number of reptile contaminant-related papers published between 1996 and 2008 by taxonomic group.......................................................................................8

Figure 1.10

A comparison of the number of contaminant- and disease-related papers published annually for reptiles from 1996 to 2008................................................8

Figure 2.1

Percentage of amphibian species with extant populations in nature in each IUCN threat category for which 6 categories of threats are believed to be operating . ............................................................................................................ 17

Figure 3.1

Status of the major lineages of reptiles according to the World Conservation Union (IUCN) Red List in 2009.................................................... 57

Figure 5.1

The conceptual model of energy and material flow provides a physiological energetics framework for evaluating exposure with traditional food chain models............................................................................. 108

Figure 10.1

Cladistic classification of “reptiles” based upon monophyletic groupings .......290

Figure 10.2

Biphenyl and polychlorinated biphenyls; both non-ortho (PCB 126) and ortho (PCBs 153 and 187) chlorinated biphenyls are described........................ 298

Figure 10.3

Polychlorinated dibenzo-p-dioxins and furans; the most toxic forms (2,3,7,8-chlorine-substituted) congeners are displayed......................................300

Figure 10.4

Examples of the relative proportion of body burdens of organochlorine pesticides (DDE, chlordane), PCBs, and PBDEs in a variety of animals, including watersnakes, turtles, and alligators....................................................300

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List of Figures

Figure 10.5

Two organochlorine pesticides, p,p’-DDE and chlordane, which are among the most common OC pesticides detected in biota............................................ 303

Figure 10.6

Three examples of polycyclic aromatic hydrocarbons often found in the environment........................................................................................................308

Figure 10.7

The surfactants perfluorooctane sulfonate and perfluorooctanoic acid.............309

Figure 10.8

Polybrominated diphenyl ethers; BDE 99 is typically found in penta formulations, whereas BDE 209 is found in the deca formulation.................... 310

Figure 10.9

Hydrocarbons found in crude oil mixtures and in some petroleum products; asphaltene is highly variable ............................................................. 312

Figure 10.10

Oral LD50s of selected organic contaminants to rats; the lower the value, the greater the toxicity........................................................................................ 314

Figure 12.1

Periodic table of elements showing the metallic elements included in the present review, and their classification according to their biological and ecotoxicological relevance.................................................................................. 339

Figure 12.2

Chronology of the cumulative number of publications dealing with the ecotoxicology of metals in reptiles..................................................................... 341

Figure 12.3

Distribution of 109Cd among reproductive tissues in female painted turtles, Chrysemys picta, 6, 24, and 192 hours after intravascular injection................. 387

Figure 12.4

Distribution of mercury, cadmium, manganese, copper, iron, and zinc among oviductal egg compartments in loggerhead turtles, Caretta caretta...... 388

Figure 12.5

Distribution of cadmium, copper, and mercury among reproductive and nonreproductive tissues of female loggerhead turtle, Caretta caretta............... 390

Figure 12.6

Distribution of cadmium among tissues of American alligators, Alligator mississippiensis, 10 days after cadmium administration (single intracardiac injection of CdCl2, 1.0 mg/kg body mass) .................................... 394

Figure 12.7

Mean mercury concentrations in tissues of wild and farm-raised American alligators, Alligator mississippiensis.................................................................. 394

Figure 12.8

Mean lead concentrations in tissues of Chelydra serpentina in the Old Lead Belt region (Missouri)............................................................................... 396

Figure 12.9

Distribution of information on metal concentrations in reptiles as indicated by the numbers of publications per reptile order and suborder, and metal........403

Figure 13.1

Sunburn and lesions on a juvenile Ambystoma tigrinum exposed to low-level irradiance in the laboratory................................................................. 451

Figure 13.2

Absorbance scans for tannic acid, tea, and water from oak leaves soaked for 2 months........................................................................................................ 455

Figure 13.3

Decreases in dissolved organic carbon increases UV-B exposure in the water column in sites in Minnesota.................................................................... 455

Figure 13.4

(A) Average oviposition depth for Ambystoma gracile egg masses at 3 sites within Mt. Rainier National Park, Washington. (B) Dissolved organic carbon concentration of water from 3 sites within Mt. Rainier National Park, Washington............................................................................................... 459

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List of Figures

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Figure 13.5

Absorption spectrum of melanin . ..................................................................... 461

Figure 13.6

Skin melanization of Bufo boreas, Bufo woodhousii, Hyla chrysocelis, and Ambystoma tigrinum showing dense melanin layers beneath the skin with variable amounts in the epidermal cells............................................................. 461

Figure 13.7

Egg jelly absorbance scans for A. gracile and R. clamitans.............................. 463

Figure 13.8

Absorbance scan for the yellow-spotted salamander......................................... 463

Figure 13.9

Egg jelly absorbance scans for Ambystoma tigrinum, Bufo boreas, and Bufo woodhousii.................................................................................................464

Figure 13.10

Absorbance spectra of Ambystoma gracile egg mass jelly as measured by a spectrophotometer............................................................................................464

Figure 14.1

The growth of amphibian toxicology papers from 1993 to 2006....................... 476

Figure 14.2

Simplified food webs used in mesocosm experiments examining larval amphibians embedded into aquatic communities.............................................. 481

Figure 16.1

Abnormalities in wild-caught North American frogs.........................................513

Figure 16.2

The digenetic trematode Ribeiroia ondatrae causes severe limb malformations in amphibians . .......................................................................... 515

Figure 16.3

Aquatic predators capable of causing limb abnormalities....................................518

Figure 16.4

Geographic distribution of Ribeiroia from amphibians..................................... 527

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List of Tables Table 1.1

Status of At-Risk Amphibians and Reptiles.............................................................2

Table 2.1

Summary of the Status of Amphibians from the Global Amphibian Assessment and Its Recent Partial Update............................................................. 15

Table 2.2

Summary of Numbers of Species in Each IUCN Category Thought to Be Subject to Threats of Different Types........................................................... 17

Table 4.1

Examples of Recent Laboratory, Mesocosm, and Field Studies Using Native North American Amphibian Species............................................... 73

Table 4.2

Examples of Recent Laboratory and Field Studies with Aquatic (Freshwater) Reptiles........................................................................ 82

Table 4.3

Examples of Recent Biomarker Studies Conducted with Amphibians or Aquatic (Freshwater) Reptiles............................................... 87

Table 6.1

Review of Amphibian Ecotoxicological Studies between 2000 and 2009.......... 170

Table 7.1

LD50 Determinations in Reptiles Administered OP Insecticides, Compared with Medial LD50 Values Calculated from Various Bird Species....................... 212

Table 8.1

Examples of Surface Water Concentrations of Atrazine...................................... 226

Table 8.2

Summary of Dose–Response Studies on Amphibians and Atrazine................... 232

Table 9.1

Concentrations of Polychlorinated Biphenyls (PCBs) and Related Compounds in Amphibians Collected from Field Studies...................................264

Table 9.2

Concentrations and Biological Concentration Factors of Organochlorine Pesticides in Amphibians Collected in the Field.................................................. 276

Table 10.1

Octanol-Water Coefficients for a Variety of Organic Contaminants................... 296

Table 10.2

Trophic Position of Selected Turtles, Snakes, Lizards, and Alligators................ 299

Table 11.1

Issues Limiting Application of Many Prior Studies of Effects of Metals on Amphibians to Ecological Questions.............................................. 328

Table 12.1

Distribution of Information of the Contamination of Reptiles with Metals According to Species, and Their Biogeographical and Ecosystem Distribution.................................................................................. 343

Table 12.2

Distribution of Information on Metals in Reproductive and Early Life Stage Compartments of Reptiles.................................................................................... 350

Table 12.3

Concentrations of Cadmium, Lead, and Mercury in Reproductive and Early Life Stage Compartments of Free-Ranging Reptiles in Comparison with Environmental Water Quality Goals.................................................................... 356

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List of Tables

Table 12.4

Concentrations of the Metals of Highest-Priority Concern (Cadmium, Lead, and Mercury) in Target Compartments (Liver, Kidney, Bone) and Potential Alternative Monitoring Compartments (Muscle, Blood, Integument, Tail Tips) of Free-Ranging Reptiles Based on Minimum, Maximum, and Mean Values Reported in the Literature per Case (i.e., per Metal, Species, Compartment, and Study)..................................................................................... 367

Table 12.5

Effects of Metals in Reptiles According to Experimental Studies (Included Are All Studies That Directly Manipulated Exposure to Metals, or to Metal-Containing Substrates).....................................................406

Table 12.6

Effects Associated with Metal Exposure in Reptiles, According to Observational Studies....................................................................................... 413

Table 12.7

Potential Adverse Health Effects Caused by Metals of Priority Concern Detected in Tissues of Reptiles............................................................................ 438

Table 13.1

Mean Wet Weight, Total Body Length (SD) for Ambystoma tigrinum Larvae from A) a 2553 M Elevation Pond (Mud Lake) and B) a 1583 M Elevation Pond (Limon Pond) Exposed to 2 Simulated Solar UV–B Intensities for 28 Days.......................................................................................... 452

Table 13.2a

Albedo of Soil Covers.......................................................................................... 453

Table 13.2b

Albedo of Vegetative Covers................................................................................ 453

Table 13.2c

Albedo of Natural Surfaces.................................................................................. 454

Table 13.3

UV Irradiance Levels and Corresponding in Amphibian Habitats Listed by Species and Altitude............................................................................................. 457

Table 15.1

List of Specific Chemical Contaminants, Abbreviations, and Basic Physical Chemical Properties.............................................................. 489

Table 15.2

Results of Toxicological Studies of Amphibian Species with Selected Emerging Contaminants Arranged by Contaminant Class........... 492

Table A.1

Aluminum (Al)..................................................................................................... 555

Table A.2

Antimony (Sb)...................................................................................................... 561

Table A.3

Arsenic (As) ........................................................................................................564

Table A.4

Barium (Ba) ........................................................................................................ 585

Table A.5

Beryllium (Be)...................................................................................................... 590

Table A.6

Cadmium (Cd)...................................................................................................... 591

Table A.7

Caesium (Cs) ........................................................................................................ 627

Table A.8

Chromium (Cr)..................................................................................................... 630

Table A.9

Cobalt (Co) ........................................................................................................644

Table A.10

Copper (Cu) ........................................................................................................ 652

Table A.11

Iron (Fe)................................................................................................................ 682

Table A.12

Tin (Sn) ................................................................................................................. 699

Table A.13

Lead (Pb).............................................................................................................. 702

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List of Tables

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Table A.14

Manganese (Mn)................................................................................................... 735

Table A.15

Mercury (Hg)........................................................................................................ 752

Table A.16

Molybdenum (Mo)................................................................................................ 786

Table A.17

Nickel (Ni)............................................................................................................ 790

Table A.18

Platinum (Pt) ........................................................................................................804

Table A.19

Rubidium (Rb)......................................................................................................805

Table A.20

Selenium (Se).......................................................................................................809

Table A.21

Silver (Ag) ............................................................................................................ 834

Table A.22

Strontium (Sr)....................................................................................................... 836

Table A.23

Thallium (Tl)........................................................................................................846

Table A.24

Titanium (Ti)........................................................................................................ 849

Table A.25

Vanadium (V)....................................................................................................... 850

Table A.26

Zinc (Zn) .............................................................................................................. 857

Table A.27

Zirconium (Zr)...................................................................................................... 885

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About the Editors Donald W. Sparling received his bachelor’s and master’s degrees in Zoology at Southern Illinois University, and his Ph.D. in biology from the University of North Dakota. After a postdoc and a brief stint in academia he began a career with the US Department of Interior, starting with the US Fish and Wildlife Service (USFWS) in 1982 and ending with early retirement from the US Geological Survey (USGS) in 2003. For most of that time, he was a research wildlife biologist at Patuxent Wildlife Research Center and conducted research on a variety of contaminants and their effects on birds and amphibians. In 2004 he returned to Southern Illinois University as associate director of the Cooperative Wildlife Research Laboratory, where he has continued contaminant research and supervised research on upland game birds. Don has approximately 100 publications in the scientific literature and has coedited 4 books, including being lead editor on the 1st edition of Ecotoxicology of Amphibians and Reptiles. Greg Linder coauthored a chapter in and helped edit the 1st edition of Ecotoxicology of Amphibians and Reptiles. During that same period, he worked with Don Sparling and Sherry Krest to convene the SETAC/Johnson Foundation workshop on multiple stressors and declining amphibian populations. For more than 30 years, Greg has worked for local, state, or federal regulatory and natural resource agencies as an applied ecologist and environmental toxicologist. Greg has worked with a wide range of plants, animals, and microorganisms, contributing to the development of toxicity test methods for fishes and aquatic invertebrates, wild mammals, amphibians, vascular plants and mosses, and a number of soil invertebrates, bacteria, and fungi. Greg works for the USGS, Columbia Environmental Research Center from his HeronWorks field office in Oregon’s Willamette Valley, where he designs and implements integrated field and laboratory studies across a wide spectrum of habitats, most often in support of ecological risk assessments and habitat restora-­ tion projects.

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About the Editors

Christine A. Bishop is a research scientist with the Canadian Federal Department of the Environment and an adjunct professor at Simon Fraser University and at the University of British Columbia. She studies the effects of habitat loss and habitat quality on wildlife. Her research focuses on the effects of multiple stressors on amphibian, reptilian, and avian populations and the recovery of populations of Species at Risk. Her doctoral studies at McMaster University examined the effects of pesticide use on birds nesting in apple orchards. Her master’s degree research at York University was the study of the effects of organochlorine contaminants on common snapping turtles (Chelydra serpentina serpentina). She received her bachelor’s degree in Agricultural Science (Honors) at the University of Guelph. During the past 26 years, she has combined her research interests with many on-the-ground conservation projects involving habitat restoration and preservation in Ontario and British Columbia, Canada. She cofounded the Canadian Amphibian and Reptile Conservation Network, has published more than 70 peer-reviewed scientific articles, and has coedited books including Ecology, Conservation, and Status of Reptiles in Canada and the 1st edition of Ecotoxicology of Amphibians and Reptiles. Sherry Krest is the Environmental Contaminants Team Leader for the Chesapeake Bay Field Office in Annapolis, Maryland, USA. Sherry obtained a B.S. in Secondary Education from Bloomsburg University and an M.S. in Environmental Biology from Hood College. She has worked for the US Army Biomedical Research and Developmental Laboratory, where she conducted her master’s study on the effects of lead on developing amphibians using the Frog Embryo Teratogensis Assay — Xenopus (FETAX). She began her career for the USFWS as an assistant course leader at the National Conservation Training Center, and then as a staff biologist in the Division of Environmental Quality in Arlington, Virginia. Today, she concentrates her efforts on addressing the use of frogs as a biomonitoring tool following the remediation of trap and skeet ranges. She is also the Natural Resource Damage Assessment and Restoration biologist and is actively involved in several cases in Maryland and Delaware.

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List of Contributors Ross A. Alford School of Marine and Tropical Biology James Cook University Townsville, Queensland, Australia Larissa L. Bailey Department of Fish, Wildlife and Conservation Biology Colorado State University Fort Collins, Colorado, USA Joseph R. Bidwell Oklahoma State University Stillwell, Oklahoma, USA Christine A. Bishop Environment Canada Science and Technology Branch Delta, British Columbia, Canada Robin D. Calfee US Geological Survey Columbia Environmental Research Center Columbia, Missouri, USA Shane R. De Solla Environment Canada Canada Centre for Inland Waters Burlington, Ontario, Canada J. Whitfield Gibbons University of Georgia Savannah River Ecology Lab Aiken, South Carolina, USA Britta Grillitsch Department of Biomedical Sciences University of Veterinary Medicine of Vienna Vienna, Austria Paula F. P. Henry US Geological Survey Patuxent Wildlife Research Center Beltsville, Maryland, USA

W. A. Hopkins Department of Fisheries and Wildlife Sciences Virginia Polytechnic Institute and State University Blacksburg, Virginia, USA Pieter T. J. Johnson Department of Ecology and Evolutionary Biology University of Colorado Boulder, Colorado, USA Sherry K. Krest US Fish and Wildlife Service Chesapeake Bay Field Office Annapolis, Maryland, USA Christine M. Lehman ABC Laboratories, Inc Columbia, Missouri, USA Greg Linder US Geological Survey Columbia Environmental Research Center Columbia, Missouri, USA Edward E. Little US Geological Survey Columbia Environmental Research Center Columbia, Missouri, USA Marc J. Mazerolle Département des sciences appliquées Université du Québec en Abitibi-Témiscamingue Rouyn-Noranda, Québec, Canada Laura L. McConnell US Department of Agriculture Beltsville Agricultural Research Center Beltsville, Maryland, USA Tana V. McDaniel Canadian Wildlife Service Burlington, Ontario, Canada

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Stacey Money Environment Canada Gatineau, Quebec, Canada Brent D. Palmer Department of Biology University of Kentucky Lexington, Kentucky, USA Bruce D. Pauli Environment Canada National Wildlife Research Centre Ottawa, Ontario, Canada Alfred E. Pinkney US Fish & Wildlife Service Chesapeake Bay Field Office Annapolis, Maryland, USA Mari K. Reeves US Fish & Wildlife Service Anchorage Fisheries & Ecological Services Office Anchorage, Alaska, USA Rick A. Relyea Department of Biological Sciences University of Pittsburgh Pittsburgh, Pennsylvania, USA

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List of Contributors

Christopher L. Rowe University of Maryland Center for Environmental Science Chesapeake Biological Laboratory Solomons, Maryland, USA Luis Schiesari Environmental Management School of Arts, Sciences and Humanities University of São Paulo São Paolo, Brazil Donald W. Sparling Cooperative Wildlife Research Laboratory Southern Illinois University Carbondale, Illinois, USA Brian D. Todd University of Georgia Savannah River Ecology Lab Aiken, South Carolina, USA Bethany K. Williams University of Missouri Division of Biological Sciences Columbia, Missouri, USA John D. Willson University of Georgia Savannah River Ecology Lab

Aiken, South Carolina, USA

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Acknowledgments As with any significant effort, there are numerous people that deserve acknowledgment for their efforts. The editors would especially like to thank the authors of the various chapters for their hard and diligent work. We would also like to thank the group of reviewers for their time in carefully examining each of the chapters and providing excellent suggestions for their improvements. Douglas J. Fort, Jamie Bacon, Chris Hamilton, and Scott McMurry took on the huge task of reviewing the entire book, and coordinating editor of SETAC Books Joe Gorsuch assisted as well. Our gratitude also goes to the Society of Environmental Toxicology and Chemistry (SETAC) and to Taylor & Francis for agreeing to publish this book, and especially to Mimi Meredith and Daniel Hatcher for careful attention to its production.

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Preface for Second Edition Ten years ago in the preface of the first edition of this book, we declared that the amount of information available on the ecotoxicology of amphibians paled in comparison to that available on birds, mammals and especially fish. Over the past 10 years, great advances have been made in understanding the effects of contaminants on amphibians. There have been more scientific, peer-reviewed papers on this topic written since 2000 than in the 30 years preceding that time. In addition to developing a better understanding of the dose-response relationships to contaminants in all the familiar chemical classes (metals, non-halogenated pesticides, organochlorinated pesticides, other halogenated organics and polyaromatic hydrocarbons), the number of species studied has increased and research is extending into emerging chemicals such as surfactants and pharmaceuticals. Perhaps the most exciting research, however, is occurring in the study of chemical interactions with other ecological stressors such as competition, predation and diseases. Another area of growth is the use of mesocosms to study the effects of contaminants on manmade communities; this type of research is revealing some surprising results compared to single organism, laboratory studies. The most critical concern at this time is maintaining the interest and momentum of amphibian-based ecotoxicological studies. In the preface of the first edition, we also lamented the paucity of studies on reptiles. Since then, there have been several studies published. In particular, prior to 2000 the vast majority of research focused on body burdens of metals and persistent organic pollutants in a few species of reptiles such as turtles. Over the past decade, there has been a shift of interest to documenting the effects of contaminants, including lethality and sublethal maladies, on this vertebrate class. However, the overall production of new information in reptilian ecotoxicology continues to be well behind that of other vertebrates. Reptiles are more than “featherless birds” — they live in different habitats, have major physiological differences and process chemicals in ways that other vertebrates do not. Thus there is a clear need to increase scientific focus on the ecotoxicology of reptiles. In 2000, we also stated that amphibian population declines were of substantial concern to the conservation community. Various causes for their declines including “fungal disease, habitat degradation, introduced predators and competitors, ultraviolet radiation, and contaminants” (p. xiv). Declining amphibian populations are still of concern although it seems that the public perception of these declines has waned. The same litany of possible causes is cited with perhaps a greater emphasis on fungal diseases, especially Batrachochytrium dendrobatidis, the cause of chytridiomycosis. We are not aware of any “smoking guns” involving contaminants as a clear and solitary cause of amphibian declines but their potential influence, especially through debilitating sublethal mechanisms remain just as real as they did 10 years ago. In fact, in light of the multitude of sublethal effects caused by contaminants — as espoused in the chapters of this book — support for contaminants as hidden and insidious causes of amphibian declines has increased substantially. In 2000 very little was known about the status of reptile species around the world. Since then, the IUCN undertook a serious examination of reptiles and concluded that a great many species are in dire straits. Undoubtedly scientists will find that contaminants have played and continue to affect declining species of reptiles. This book is intended to provide a current synthesis of the scientific state of amphibian and reptile ecotoxicology. We have updated many of the chapters in the first edition, dropped a few along the way, and included a few more to present topical issues. In preparing this book, we sought

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xxiv

Preface for Second Edition

out many of the same authors as in the first edition and included several others that have made important contributions to our understanding of amphibian and reptile ecotoxicology over the past decade. The choice of authors is always a difficult one because for each one that is invited, there are several others whom you cannot invite. As with the first edition, all praise should go to the authors who have contributed to this book and any complaints can be directed to the editors. – Donald W. Sparling – Greg Linder – Christine A. Bishop – Sherry K. Krest

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Preface from the First Edition An international concern for the status and welfare of amphibians and their populations has been building since the late 1980s, reaching a turning point in the mid-1990s when two independent sets of events occurred. One set was several conferences that were held across the globe to discuss the status of amphibian populations. For several years prior to these conferences, similar discussions were stymied from reaching definitive conclusions by a lack of long-term population monitoring and historical data. At that time, however, scientists generally came to agree that the apparent declines were genuine and that numerous populations and species were at risk or even extirpated. Several reasons for these declines were espoused, including fungal disease, habitat degradation, introduced predators and competitors, ultraviolet B radiation (with and without interaction by chemicals), and contaminants. At the present time, none of these hypotheses has been universally supported. The other significant event in the mid-1990s was the observation of numerous malformed frogs in Minnesota. Again, a lack of readily available historical data resulted in more questions than answers on the significance of this observation. As a result, several extensive surveys have been implemented in the US and Canada to determine the extent of these malformations and to identify possible causes. Proposed causes include ultraviolet B radiation, parasitism, and contaminants. In an attempt to further understand the possible role of contaminants in amphibian population declines and malformations and in order to develop research that would help address these problems, each of us independently began a review of the existing literature. As we did so, we quickly became aware that information on contaminant effects and burdens in amphibians was extremely scarce and dispersed compared to that of other vertebrates, especially fishes, birds, and mammals. Even more apparent was an almost total lack of knowledge about contaminant effects on reptiles. Some body-burden data could be found, but hardly anything at all was found on the effects of these contaminants on the health or survival of reptiles. What information existed was heavily skewed towards a few species of turtles and clearly was not representative of the class. In response to the absence of a concerted effort to evaluate the effects of contaminants on amphibians and reptiles, we endeavored to enlist the efforts of other researchers and develop a current state of science and synthesis of what is known with the hopes that such a compilation would spur additional inquiry and research. The results of these efforts follow. In developing the book, we realized that several audiences might have an interest in its content. However, two general groups were foremost in our minds. First were the herpetologists, ecologists, and zoologists who might be interested in amphibians and reptiles in their own right but who might not have a strong background in ecotoxicology. The other major group was ecotoxicologists, resource managers, and policymakers who are versed in contaminant ecology and want to know more about amphibians and reptiles and how they compare to the better known classes of vertebrates. To meet the needs of both groups we have arranged for a variety of chapters covering 1) basic ecology, distribution, and physiology of these vertebrates; 2) syntheses of the existing information on specific groups of contaminants and herpetofauna; and 3) issues of risk assessment and study designs for those wishing to conduct additional research. Through this whole process and praise should be given to the authors, and any complaints can be directed toward us. – D. Sparling – G. Linder – C. Bishop xxv

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1

Recent Advancements in Amphibian and Reptile Ecotoxicology Donald W. Sparling, Greg Linder, Christine A. Bishop, and Sherry K. Krest

Contents 1.1 Current Status of Ecotoxicological Research on Amphibians and Reptiles.............................2 1.1.1 Source of Publication Information................................................................................2 1.1.2 Comparisons among Vertebrates...................................................................................3 1.1.3 Productivity within Amphibians and Reptiles..............................................................3 1.1.3.1 Amphibians.....................................................................................................5 1.1.3.2 Reptiles...........................................................................................................7 1.1.4 Conclusions....................................................................................................................9 1.2 What’s in This Book?................................................................................................................9 References......................................................................................................................................... 10

When the first edition of Ecotoxicology of Amphibians and Reptiles was published in 2000, I reviewed the state of the literature from 1972 through 1998 (Sparling et al. 2000). That review covered 11 271 contaminant citations listed in Wildlife Review and Sports Fisheries Abstracts published by the US Fish and Wildlife Service. Among its findings, only 2.7% of the cited papers were on amphibians and 1.4% on reptiles. This equated to an average annual rate of 11.5 citations for amphibians and 6 for reptiles, although the distribution of citations was not homogeneous through the years. In contrast, 61.8%, or 280, contaminant citations per year were on fish. Among the amphibian citations, most focused on effects, 23% dealt with metals, 22% with acid precipitation, and 19% with nonchlorinated pesticides. The remaining 38% covered all of the other contaminants of interest at that time. Almost all of the citations on reptiles dealt exclusively with residues, and turtles (Chelonia) were overrepresented compared to the percent of reptilian species comprised by this order. The most important categories of contaminants included metals (24%), organochlorine pesticides (23%), and polychlorinated biphenyls (PCBs) (19%), all persistent pollutants. At that time we raised the question why these 2 classes of vertebrates should be so underrepresented in the contaminant literature. Amphibians make up approximately 20% and reptiles 28% of known vertebrate species, so the number of contaminant citations was far fewer than would be predicted by species richness. In addition, both classes are extremely important ecologically in their respective habitats (Stebbins and Cohen 1995). More recently, a series of papers (Ranvestal et al. 2004; Regester et al. 2006; Regester and Whiles 2006; Whiles et al. 2006) have documented that amphibian declines may very seriously and negatively affect the nutrient balance, species diversity and richness, and energy flow of ponds, streams, and associated uplands. 1

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Ecotoxicology of Amphibians and Reptiles

Table 1.1 Status of At-Risk Amphibians and Reptiles Amphibians Number of species evaluated Critically rare or endangered Endangered Vulnerable Total (percent of evaluated)

5918 442 738 620 1800 (30.4)

Reptiles 664 73 100 125 298 (44.9)

Source: IUCN (2007).

Many species of amphibians and reptiles around the world are endangered or threatened with endangerment. Of 5915 amphibian species evaluated, the IUCN (2007) lists 1808 (30%) at risk (Table 1.1). This contrasts with only 127 similarly listed species in 1996 (IUCN 1996). Even more telling, of approximately 6000 species of reptiles, the IUCN (2007) has evaluated the status of only 1385 (ca. 23%). Among these, 549 (39.6%) are considered at risk. To the extent that contaminants have a role in the imperiled status of amphibians and reptiles, further research is demanded (Gibbons et al. 2000; Hopkins 2000). Since the 2000 review there has been a dramatic increase in the number of research studies and papers focusing on the effects and burdens of contaminants on these 2 important classes of vertebrates. While we would like to think that the publication of the first edition of Ecotoxicology of Amphibians and Reptiles had a major role in this increased interest, other factors such as symposia (e.g., Midwest Declining Amphibian Conference, Milwaukee, Wisconsin, 1998; Ecological Society of America, Memphis, Tennessee, 2006), a temporary increase in federal funding on problems associated with amphibian declines and monitoring (e.g., US Geological Survey’s Amphibian Research and Monitoring Initiative and North American Reporting Center for Amphibian Malformations, US Fish and Wildlife Service’s investigation for malformed frogs on national wildlife refuges), other, nonfederal organizations (e.g., Partners in Amphibian and Reptile Conservation [PARC], Declining Amphibian Population Task Force [DAPTF], now united with International Union for the Conservation of Nature), and publications (e.g., Lannoo 1998; Linder et al. 2003a, 2003b; Gardner and Oberdörster 2006) were very instrumental in this surge of interest. This chapter provides an overview of the increased research efforts since the publication of the first edition of Ecotoxicology of Amphibians and Reptiles. We finish with a description of what readers may expect to find in the subsequent chapters of this book.

1.1 Current Status of Ecotoxicological Research on Amphibians and Reptiles 1.1.1

Source of Publication Information

Because Wildlife Review and Sports Fisheries Abstracts are no longer published, we had to rely on a different source of citations. The increased technology in literature review services since 2000 greatly expanded the opportunity to examine a large number of citations in a comparatively short period of time. However, differences in search methodology and the extensive databases that are now available affect meaningful comparisons between the first edition and this chapter. To help mitigate this problem and to ensure continuity with the 2000 edition of this book, we extended our search back to 1996 through 2008 to develop a more historical perspective on changes in amphibian and reptile studies. We used the ISI Web of Knowledge, specifically Web of Science, as our data source.

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Recent Advancements in Amphibian and Reptile Ecotoxicology

3

14000 12000

Fish Mammal Bird Amphibian Reptile

Number of Pubs

10000 8000 6000 4000 2000 0 Taxon

Figure 1.1  Total number of contaminant-related papers published between 1996 and 2008 by vertebrate class.

1.1.2  Comparisons among Vertebrates Initially, we ran a search on each of the vertebrate classes to derive a relative measure of emphasis. In this search we did not examine specific titles. We matched the keywords metal*, organochlorine, pesticide, PCB*, PAH, contaminant, and toxic*, joined by the Boolean operator or, with each of the taxa and determined the number of citations. The asterisk (*) is a wild card or generic expression allowing any subsequent characters through if the preceding characters are matched. For mammals, we made the additional choice of excluding citations with keywords human, mouse, and rat to reduce the number of citations that were primarily or exclusively oriented to human health. The choice of keywords in literature searches is a subtle science, and we acknowledge that our search was not exhaustive. However, we feel that this method provided a reasonable comparison of research productivity by vertebrate class. By far, the number of citations for fish remained much higher than those of other classes (Figure 1.1). Of the 17 375 citations examined, 11 601 (66.7%) were for fish, 3457 (19.9%) for mammals, 1520 (8.8 %) for birds, 645 (3.8%) for amphibians, and 152 (0.8%) for reptiles. This distribution is very similar to the one found in 2000, except that the relative position of birds and mammals is switched. Only a very small fraction of the literature on vertebrate ecotoxicology pertains to amphibians or reptiles.

1.1.3  Productivity within Amphibians and Reptiles In a more refined search, we examined each year (1996 to 2008) and entered the keywords amphibian*, salamander, frog, toad, newt, Bufo, Rana, Hyla, Acris, Pseudacris, and Ambystoma for amphibians and reptile*, alligator, snake, turtle, lizard, Chelydra, and anole for reptiles. In addition, for reptiles we excluded venom*, grass, and river because of frequent noncontaminant hits on these keywords. Each title was reviewed to determine if the citation was contaminant related and to which class of compounds and taxonomic group it could be attributed. In scoring, a given citation may have had 1 or more “hits,” with a hit consisting of a specific class of contaminant, taxon, or analysis. For example, a citation with the hypothetical title “Effects of Heavy Metals and PCBs on Ambystoma gracile and Bufo cognatus” would have received 1 hit each for effects, heavy metals, PCBs, Ambystoma, and Bufo. Abstracts and publications were searched when titles alone were insufficient to identify types of contaminants examined. In all, 25 998 amphibian and 15 057 reptile citations were searched. We recognized the following categories or classes of contaminants:

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Ecotoxicology of Amphibians and Reptiles

1) pesticides — including all nonhalogenated pesticides; 2) metals — particularly heavy metals such as lead, arsenic, cadmium, zinc, copper, and mercury, but could also include metalloids and other metals, such as aluminum, that can be toxic under certain conditions; 3) ultraviolet radiation — including UVA and UVB, whose environmental levels have been accentuated in recent times; 4) simazines such as atrazine; 5) nitrogenous compounds, particularly nitrates, nitrites, and ammonia, that are used as fertilizers; 6) PCBs, dioxins, and furans; 7) polycyclic aromatic hydrocarbons (PAHs); 8) organochlorinated pesticides such as endosulfan, dichlorodiphenyltrichloroethane (DDT), and dichlorodiphenyltrichloroethylene (DDE); and 9) other — including polyhalogenated diphenyl ethers, ingested plastics in the case of sea turtles, radioactive molecules, pharmaceuticals, and various biocides and antiseptics.

There was a significantly increasing trend in the total number of contaminant and noncontaminant citations by year for reptiles (r = 0.9240, p < 0.0001) but not for amphibians (r = 0.2059, p = 0.4997) (Figure 1.2), suggesting that there has been an overall increase in productivity in reptile biology. However, amphibians are used extensively in a variety of different research contexts, including embryology, physiology, genetics, and medical and human health studies that vary independently of ecological or contaminant-related interests; thus, large volumes of research are continually being produced for this class. Reptiles are used less in these types of studies, and their increase in the number of citations through time may more closely mirror greater interest in reptile behavior, ecology, and conservation status. Annual numbers of contaminant-related amphibian papers, although showing year-to-year variation, significantly increased from 1996 to 2008 (r = 0.8566, p = 0.0002) (Figure 1.3), as did their relative contribution to the total number of amphibian papers published (r = 0.7507, p = 0.0031) (Figure 1.4). However, the percent of total publications that are contaminant-related has remained relatively stable since 2002 (Figure 1.4). The total number of contaminant-related publications between 2000 and 2005 exceeded that for the combined 22 years prior to 2000 for amphibians. In contrast, the number of contaminant-related papers for reptiles has been more erratic than for amphibians but has shown a significant increasing trend through the years (r = 0.7098, p = 0.0066) (Figure 1.3). Similarly, the percent of total publications that are contaminant related has not significantly increased since 1997 (r = 0.3873, p = 0.1910) (Figure 1.4). These data indicate

Number of Pubs

3600 3200 2800 2400 2000 1600 1200 19 96 19 97 19 98 19 99 20 00 20 01 20 02 20 03 20 04 20 05 20 06 20 07 20 08

800

Year

Figure 1.2  Annual numbers of scientific publications for amphibians and reptiles between 1996 and 2008 as listed in ISI Web of Science.

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5

Number of Pubs

90 75 60 45 30 15 0

96 997 998 999 000 001 002 003 004 005 006 007 008 2 2 2 1 1 1 2 2 2 2 2 2

19

Year

Figure 1.3  Number of contaminant-related papers published for amphibians and reptiles between 1996 and 2008. (Data from ISI Web of Science.)

that increased interest in amphibian ecology since worldwide population declines were generally accepted has resulted in elevated scientific productivity. By comparison, reptile ecotoxicology remains relatively unexplored and unknown. From 2006 through 2008 only 17, 20, and 21 citations, respectively, could be found for any contaminant study in reptiles, whether it focused on effects or residues. As Hopkins (2000) pointed out, this dearth of knowledge should be of great concern. 1.1.3.1  Amphibians Among amphibians, the most intensely studied class of contaminants has been nonchlorinated pesticides, including organophosphorus, carbamate, and pyrethroid compounds (Figure 1.5). This class is followed in order by metals; ultraviolet radiation; endocrine disruptors (which include several classes of contaminants); PCBs and related molecules; simazines (mostly atrazine, of course); nitrogenous compounds, including nitrates, nitrites, and ammonia; acidity; polycyclic aromatic hydrocarbons; and organochlorine pesticides. Changes in emphasis from pre-2000 include a sharp decrease in publications on acid precipitation and an increased focus on pesticides, ultraviolet radiation, endocrine disruption, and atrazine. Slightly less than 6% of the contaminant papers published between 1996 and 2008 dealt extensively with residue data without providing new information on the effects of those residues. 5

Percent

4 3 2 1 0

96 997 998 999 000 001 002 003 004 005 006 007 008 1 1 1 2 2 2 2 2 2 2 2 2

19

Year

Figure 1.4  Percent of all scientific publications for amphibians and reptiles between 1996 and 2008 that were contaminant related. (Data from ISI Web of Science.)

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Ecotoxicology of Amphibians and Reptiles

Number of Pubs

250 200 150 100 50 0 Contaminant Class

Pesticides Metals Other Ultraviolet Endocrine PCB/PCDD Simazines Nitrogenous Acidity PAH Organochlorine

Figure 1.5  Contaminant-related papers published on amphibians between 1996 and 2008 by chemical class. (Data from ISI Web of Science.)

Number of Pubs

Ranids have been the most studied family of amphibians (Figure 1.6), accounting for more studies than Bufonidae, Hylidae, Ambystomatidae, and other families combined. For Xenopus, we only counted studies that had direct ecotoxicological relevance. Even with that constraint, Xenopus laevis remains the most commonly studied amphibian species. In comparison to other issues related to amphibian population declines, contaminant ecology remained the leading topic of study up to 2007. From 2005 there has been a surge in disease-related papers, no doubt stimulated by findings that chytrid fungus (Batrachochytrium dendrobatidis) and ranaviruses have been linked to population declines (Longcore et al. 1999; Daszak et al. 1999) (Figure 1.7). In 2007 the number of disease-associated papers exceeded the number of contaminantrelated papers for the first time in at least a decade, and the number of disease papers continued to surpass the number of contaminant papers in 2008. Extremely few papers were published on malformations prior to 1996, but the topic became important after the identification of several malformed tadpoles in Minnesota in 1995; this increase lasted about 10 years and then dissipated. Climate change, which has been cited as a possible mechanism of amphibian population decline (e.g., Davidson et al. 2001; Pounds 2001; Carey and Alexander 2003; Daszak et al. 2005), has not received rigorous investigation. The scientific community is just beginning to explore possible interactions between contaminants in the broad sense and other stressors. We identified 36 research papers published since 1996 with clearly specified objectives of testing the interaction between one or more contaminants (including 360 330 300 270 240 210 180 150 120 90 60 30 0

Randiae Xenopus Bufonidae Hylidae Ambystomatidae Salamandaridae Australian Taxa Plethodontidae Other

Taxa

Figure 1.6  Contaminant-related papers for amphibians that were published between 1996 and 2008 by taxonomic group. (Data from ISI Web of Science.)

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Recent Advancements in Amphibian and Reptile Ecotoxicology

Number of Pubs

105 90 75 60 45 30 15 0

Contaminants Diseases Climate Malform Declines

96

19

98

19

00

20

02

20

04

20

06

20

08

20

Year

Figure 1.7  A comparison of the numbers of papers published for amphibians between 1996 and 2008 by associated stressor. (Data from ISI Web of Science.)

ultraviolet radiation) and other stressors. Of these, 19 (53%) investigated the interactive effects of predation, 12 (33%) examined disease, and 5 (14%) looked at interspecific competition or other factors. Relyea’s chapter in this book (Chapter 14) comprehensively examines the interaction between contaminants and other stressors. 1.1.3.2  Reptiles In reptiles, the most studied class of contaminants is heavy metals, followed by endocrine disrupting chemicals, organochlorine pesticides, and nonchlorinated pesticides (Figure 1.8). Simazines, ultraviolet radiation, acidification, and nitrates/nitrites were lightly studied. However, compared to the years prior to 2000, even a few papers marks increased interest in these topics. Metals and organochlorine pesticides were frequently cited throughout both time periods, but more papers were published on metals and pesticides in the years between 2000 and 2008 than in the combined 28 years preceding 2000. Whereas the vast majority of papers written prior to 2000 were based on residue information with little data on effects, 64% of the papers published after 2000 focused on effects. Thus, we are in the beginning stages of understanding how contaminants affect reptiles. By far, the most intensively studied reptilian species in ecotoxicology is the American alligator (Alligator mississippiensis); (Figure  1.9). Largely due to the research of Guillette and others 80

Metals

Number of Pubs

70

Endocrine

60

Organochlorine PCB/PCDD

50

Pesticides

40

Other PAH

30

Simazines

20

Nitrogenous UVB

10

Acidity

0 Contaminant Class

Figure 1.8  Number of contaminant-related papers published between 1996 and 2008 on reptiles by type of contaminant. (Data from ISI Web of Science.)

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8

Ecotoxicology of Amphibians and Reptiles 60 Number of Pubs

50 40 30 20 10 0

FW Turtle Crocodylia Sea Turtle

Squamata

Serpentes

Taxa

Figure 1.9  Number of reptile contaminant-related papers published between 1996 and 2008 by taxonomic group. (Data from ISI Web of Science.)

(Jagoe et al. 1998; Guillette et al. 1999, 2000), alligators contribute the majority of studies under Crocodylia, and many of these papers have been on endocrine disruption and persistent organic pollutants. Freshwater turtles, especially common snapping turtles (Chelydra serpentina), would be next in importance, followed by sea turtles and Squamata (lizards). Contaminant papers on Serpentes (snakes) are uncommon. The preference for snapping turtles is probably due to their large size, relatively high visibility and capture rates, widespread distribution, long lifespan, and bottom dwelling; hence, they have ample opportunity to assimilate and accumulate persistent organic pollutants and metals and are easily studied. When compared to amphibians, existing literature on other factors affecting reptile declines (Gibbons et al. 2000) remains sparse. Whereas a variety of papers deal with habitat needs and use, behavior, and general ecology, only disease-related studies were specifically related to population declines, and many of these were related to captive animals. This category includes data on zoonoses, parasitism, and incidence or diagnosis of disease under laboratory, zoological, and ecological conditions. More than twice the number of papers was published on disease issues than on contaminants from 2000 through 2008 (Figure 1.10). Only a scattering of papers were found for reptile declines or climate change.

Number of Pubs

60 50 40

Contaminants

30

Disease

20 10 19 9 19 6 9 19 7 1998 9 20 9 2000 0 20 1 0 20 2 0 20 3 0 20 4 0 20 5 2006 0 20 7 08

0

Year

Figure 1.10  A comparison of the number of contaminant- and disease-related papers published annually for reptiles from 1996 to 2008. (Data from ISI Web of Science.)

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Recent Advancements in Amphibian and Reptile Ecotoxicology

9

1.1.4  Conclusions Considerable progress has been made in developing new information on the ecotoxicology of amphibians and reptiles over the past few years compared to the nearly 3 decades preceding 2000. Since 2000, total publications in both vertebrate classes have exceeded those appearing from 1972 to 2000 combined. The increased interest and productivity in ecotoxicology has been most notable in amphibians. Significant increases have occurred in contaminant ecology of amphibians in both the number of papers and the percent of all amphibian papers published. No doubt, these increases are related to the upsurge in interest due to widely accepted phenomena of population declines and species extinctions. Among reptiles, however, the increased productivity has not been as obvious, even though many species appear to be suffering population losses as great as those in amphibians (Hopkins 2000; Chapters 2 and 3, this book). Whereas the total number of contaminant-related papers increased for this class since 2000, so did the total number of papers, and there was no noticeable surge in ecotoxicology of reptiles. Moreover, the actual number of papers that are published annually on contaminants and reptiles is very meager. Using the number of citations as a guide, specific gains have been made in certain areas of amphibian and reptile ecotoxicology during the last few years. Most notably, research is being conducted and published on the effects of contaminants on reptiles, not just on body burdens. Whereas these effects papers derived from a very few species, mostly alligators, turtles, and a couple of lizards, new information is being obtained. Another significant trend is an increase in the number of amphibian species that have been examined. Whereas there are a few more intensively widely studied species, notably Xenopus laevis, Rana pipiens, R. sphenocephala, Bufo fowlerii, B. cognatus, B. bufo, Hyla versicolor, and various Ambystomids, more than 12 families have been investigated at least once. No studies were found on the poorly known order Gymnophiona (caecillians). Much more work needs to be done for a fuller understanding of contaminant effects in amphibians and reptiles. The first goal should be to obtain more information on the effects of contaminants on reptiles. For many years the only contaminant data available for this class were on body burdens or residues of persistent organics. In addition to looking forward to new scenarios, it may be useful for some laboratory studies to take a retrospective look and determine if the reported burdens could have been detrimental to individuals or populations. Additional research on the interaction between chemicals with other stressors is of nearly equal importance. Contaminants will most often negatively influence populations in subtle ways (Sparling 2003; Chapter 14, this volume) in consort with disease, predation, competition, and food availability. Bioindicator development is another area that needs to be developed further in both vertebrate classes. Temperature-dependent sex determination in many reptiles and some amphibians offers many opportunities to examine sex-related hormone disruption, and metamorphosis in anuran amphibians provides an excellent background for thyroiddisrupting chemicals. Analyses of how contaminant effects vary during entire amphibian life cycles can provide new understanding on the effects of these contaminants on population declines. It is our hope that readers of this volume will be inspired in many ways to develop new research thrusts in the ecotoxicology of amphibians and reptiles.

1.2  What’s in This Book? As with the first edition of Ecotoxicology of Amphibians and Reptiles (Sparling et al. 2000), the second edition can be divided into a few parts based on the general content. Whereas readers of the first edition will recognize the names of some authors, the editors for the second edition make a point of inviting some new authors so as to obtain a different perspective on some of the major contaminant issues affecting amphibians and reptiles. In all cases we encouraged authors to use similar chapters in the first edition as jumping off places. Authors were encouraged to

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emphasize new information published after the first edition came out but to draw from that source when appropriate. The first portion of the second edition may be thought of as background material. In it readers may find current thoughts on the status of amphibians and reptiles (Chapters 2 and 3), an overview of contaminant ecology in these 2 classes (Chapter 4), and a thoughtful perspective on amphibian and reptile physiology from a contaminants exposure and effects perspective (Chapter 5). The authors for Chapter 2 hope that their presentation will whet the appetites of readers for the rest of the book. The second portion of the book focuses on the effects and residues of specific contaminant groups on amphibians and reptiles. Chapters 6 and 7 deal with nonchlorinated pesticides such as organophosphorus, carbamate, and pyrethroid compounds. In the last few years the herbicide atrazine has been extensively studied with regard to amphibians because of its widespread and common use and because there has been a great deal of controversy over whether it interferes with normal development of gonads at very low concentrations. Christine Bishop, one of the editors of this volume, brings the reader up to date (as of late 2008) on this important chemical (Chapter 8). Organic contaminants, including chlorinated pesticides, polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs), and their related dioxins and furans, are the focus of Chapters 9 and 10. Metals and metalloids are discussed in Chapters 11 and 12. Section 3 consists of chapters that might be called “contaminants plus.” Chapter 13 reviews work conducted on amphibians (mostly) and ultraviolet radiation (UV). Although it can be argued that UV is not a contaminant in the usual sense, we have included this chapter because UV radiation has increased in recent decades, in part due to the thinning of the ozone layer, which is affected by contaminants, and because UV affects the toxicity of some contaminants, such as PAHs. The multiple stressors chapter (Chapter 14) focuses on the increasing body of knowledge on how contaminants interact with other stressors, such as disease, competition, and predation, to potentially exert greater effects than the same contaminants by themselves. The implications of these interactions for making exposure tests increasingly realistic are obvious. While we have a working knowledge of the effects of many contaminants that have been around for awhile, there is a whole new cadre of chemicals that have recently or are about to come on the market for which we have very little or no information. Some of these are potentially alarming. Chapter 15 examines some of the major chemicals in this arena. As presented in the first part of this chapter, amphibian malformations have received considerable attention since 1995. While the first edition reviewed the current literature up to 2000, Chapter 16 in this volume reports on a major study conducted since that time and reviews the work of many scientists. Over the past few years efforts led by the US Geological Survey’s Amphibian Research and Monitoring Initiative have greatly improved methods for monitoring and enumerating amphibian populations. Because many contaminant-related field studies rely on some form of surveying, Chapter 17 quickly encapsulates the most important advances for surveying and monitoring. The final chapter, or epilogue, is the editors’ almost random thoughts on what we learned during the development of this book and where we believe the science of ecotoxicology of amphibians and reptiles should be heading.

References Carey C, Alexander MA. 2003. Climate change and amphibian declines: is there a link? Divers Distrib 9:111–121. Daszak P, Berger L, Cunningham AA, Hyatt AD, Green DE, Speare R. 1999. Emerging infectious diseases and amphibian population declines. Emerg Infect Dis J Cent Dis Cont Prevent 1–21. Available from: http:// www.cdc.gov/ncidod/eid/vol5no6/daszak.htm. Daszak P, Scott DE, Kilpatrick AM, Faggioni C, Gibbons JW, Porter D. 2005. Amphibian population declines at Savannah River site are linked to climate, not chytridiomycosis. Ecology 86:3232–3237. Davidson C, Shaffer HB, Jennings MR. 2001. Declines of the California red-legged frog: climate, UV-B, habitat, and pesticides hypotheses. Ecol Appl 11:464–479.

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Gardner SC, Oberdörster E. 2006. Toxicology of reptiles. Boca Raton (FL): CRC Press, Taylor & Francis Group. Gibbons JW, Scott DE, Ryan TJ, Buhlmann KA, Tuberville TD, Metts BS, Greene JL, Mills T, Leiden Y, Poppy S, Winne CT. 2000. The global decline of reptiles, déjà vu amphibians. Bioscience 50:653–666. Guillette Jr LJ, Brock JW, Rooney AA, Woodward AR. 1999. Serum concentrations of various environmental contaminants and their relationship to sex steroid concentrations and phallus size in juvenile American alligators. Arch Environ Contam Toxicol 36:447–455. Guillette LJ, Crain DA, Gunderson MP, Kools SAE, Milnes MR, Orlando EF, Rooney AA, Woodward AR. 2000. Alligators and endocrine disrupting contaminants: a current perspective. Am Zool 40:438–452. Hopkins WA. 2000. Reptile toxicology: challenges and opportunities on the last frontier in vertebrate ecotoxicology. Environ Toxicol Chem 19:2391–2393. [IUCN] International Union for the Conservation of Nature. 1996. 1996 red list of threatened animals. IUCN, Gland, Switerland, and Cambridge, England. [IUCN] International Union for the Conservation of Nature. 2007. 2007 red list of threatened animals. Jagoe CH, Arnold-Hill B, Yamochko GM, Winger PV, Brisbin Jr IL. 1998. Mercury in alligators (Alligator mississippiensis) in the southeastern United States. Sci Total Environ 213:255–262. Lannoo M, editor. 1998. Status and conservation of midwestern amphibians. Iowa City (IA): University of Iowa Press. Linder G, Krest S, Little EE, Sparling DW. 2003b. Multiple stressor effects in relation to declining amphibian populations. STP1443. West Conshohocken (PA): ASTM International. Linder G, Krest S, Sparling DW. 2003a. Amphibian decline: an integrated analysis of multiple stressor effects. Pensacola (FL): SETAC Press. Longcore J, Peskier AP, Nichols DK. 1999. Batrachochytrium dendrobatidis gen. et sp. nov., a chytrid pathogenic to amphibians. Mycologia 91:219–227. Pounds JA. 2001. Climate and amphibian declines. Nature 410:639–640. Ranvestal, AW, Lips KR, Pringle CM, Whiles MR, Bixby RJ. 2004. Neotropical tadpoles influence stream benthos: evidence for the ecological consequences of decline in amphibian populations. Fresh Biol 49:274–285. Regester KJ, Lips KR, Whiles MR. 2006. Energy flow and subsidies associated with the complex life cycle of ambystomatid salamanders in ponds and adjacent forest in southern Illinois. Oecologia 147:303–314. Regester KJ, Whiles MR. 2006. Decomposition rates of salamander (Ambystoma maculatum) life stages and associated energy and nutrient fluxes in ponds and adjacent forest in southern Illinois. Copeia 2006:640–649. Sparling DW. 2003. A review of the role of contaminants in amphibian declines. In: Hoffman D, Rattner BA, Cairns J, editors, Handbook of ecotoxicology. Boca Raton (FL): Lewis Publishers, p 1099–1128. Sparling DW, Linder G, Bishop CA, editors. 2000. Ecotoxicology of amphibians and reptiles. Pensacola (FL): SETAC Press. Stebbins RC, Cohen NW. 1995. A natural history of amphibians. Princeton (NJ): Princeton University Press. Whiles MR, Lips KR, Pringle CM, Kilham SS, Bixby RJ, Brenes R, Connelly S, Colon-Gaud JC, Hunte-Brown M, Huryn AD, Montgomery C, Peterson S. 2006. The effects of amphibian population declines on the structure and function of Neotropical stream ecosystems. Front Ecol Environ 4:27–34.

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2

Declines and the Global Status of Amphibians Ross A. Alford

Contents 2.1 Amphibian Declines: Defining the Problem........................................................................... 13 2.2 Potential Causes of the Amphibian Decline Crisis................................................................. 16 2.2.1 Habitat Destruction...................................................................................................... 18 2.2.2 Exotic Predators...........................................................................................................20 2.2.3 Environmental Contamination....................................................................................20 2.2.4 Human Utilization....................................................................................................... 22 2.2.5 Disease......................................................................................................................... 22 2.2.5.1 Ranaviruses................................................................................................... 23 2.2.5.2 Saprolegnia and Other Fungal Pathogens....................................................24 2.2.5.3 Amphibian Chytrid Fungus..........................................................................24 2.2.6 Amphibian Limb Malformations and Their Relation to Parasitic Disease................. 29 2.2.7 Ultraviolet Radiation................................................................................................... 29 2.2.8 Climate Change........................................................................................................... 30 2.3 Management of Amphibian Populations................................................................................. 31 2.4 Summary and Conclusions...................................................................................................... 32 References......................................................................................................................................... 32 The global diversity and status of both amphibians and reptiles at the end of the 20th century were described in thorough detail in the previous edition of the current volume (McDiarmid and Mitchell 2000). During the 1990s, however, it became clear that the diversity and status of amphibians were changing rapidly, with substantial declines in amphibian populations in many regions (Alford and Richards 1999). There has also been widespread concern regarding possible increases in the prevalence of developmental abnormalities and malformations in amphibians (Johnson et al. 2003). Great advances have been made in understanding these problems since 1999. This chapter concentrates on defining the problems, illustrating and interpreting our understanding of their causes, and developing ideas regarding how the problems can be managed.

2.1  Amphibian Declines: Defining the Problem It has been almost 20 years since the problem of global amphibian declines became widely recognized (Barinaga 1990; Beebee et al. 1990; Blaustein and Wake 1990; Vitt et al. 1990). Early investigation of the problem was impeded by a lack of data on the normal behavior and population dynamics of amphibians. The extremely dynamic nature of amphibian populations (Pechmann et al. 1991; Sjogren-Gulve 1991) made it difficult to determine whether observed declines might be part of normal population processes, leading to an initial debate over whether there really was a widespread problem (Wake 1991; Crump et al. 1992; Blaustein 1994; Pechmann and Wilbur 1994; Travis 1994). Herpetologists quickly realized, however, that the great extent and precipitous manner of at 13

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least some declines (e.g., Fellers and Drost 1993; Richards et al. 1993; Vial and Saylor 1993) were not consistent with even the most extreme predictions of models of population or metapopulation dynamics. A formal test of models by Pounds et al. (1997) showed that at least some declines were inconsistent with normal population fluctuations. Many herpetologists adopted the position advocated by McCoy (1994), affirming that unusually severe declines had recently occurred, but that additional data were necessary to clarify the extent and nature of the problem. Although the phenomenon of amphibian declines was first recognized and publicized in the early 1990s, they appear to have commenced earlier. Precipitous declines of regional faunas occurred in the 1970s and 1980s in North America (Corn and Fogleman 1984; Carey 1993), Central America (Pounds and Crump 1994), South America (Heyer et al. 1988; Eterovick et al. 2005), and Australia (Czechura and Ingram 1990). Widespread declines may have commenced in the 1960s (Houlahan et al. 2000), but a reanalysis of the same data (Alford et al. 2001) suggested that widespread declines primarily commenced in the 1990s. Regardless of when they commenced, there is presently a consensus among amphibian biologists that the group as a whole has suffered and continues to suffer calamitous declines in most regions that have sufficient data available to detect them. The response to the amphibian decline crisis commenced in earnest in February 1990, when a meeting sponsored by the US National Academy of Sciences concluded that there appeared to be a genuine problem, and that it should be addressed by an international working group of scientists (Heyer and Murphy 2005). This led to the formation of the International Union for Conservation of Nature/Species Survival Commission (IUCN/SSC) Declining Amphibian Populations Task Force (DAPTF) in December 1990 (Heyer and Murphy 2005). Throughout the 1990s, until its merger with the Amphibian Specialist Group of IUCN/SSC in 2007, the DAPTF played a major role in the global coordination of efforts to understand and address the problem of amphibian declines. It established scientific working groups by discipline and by region, organized meetings and symposia to facilitate exchange of data on the problem, and facilitated the creation of the first book of standardized protocols for surveying amphibian populations (Heyer et al. 1994). It also established a seed grant program to fund start-up projects, particularly targeting projects and areas of the world that might have difficulty in obtaining funding from other bodies; as of 2002, 95 projects had been awarded almost US$210  000, resulting in numerous publications and allowing many awardees to obtain additional larger-scale funding based on their preliminary data (Heyer and Murphy 2005). Regional and national working groups associated with the DAPTF produced published reports on the status of amphibians in the Lesser Antilles, the Commonwealth of Independent States, Canada, Australia, and the United States (Campbell 1999; Heyer and Murphy 2005). These gave a clearer focus to the problem and began to explore the nature of its causes and possible solutions. A series of review papers published in the 1990s presented the issues and revealed the rapid increase in knowledge of the status of amphibians and of factors affecting that status. Early efforts (Blaustein et al. 1994b; Kuzmin 1994; Pechmann and Wilbur 1994) emphasized the lack of knowledge of, and need for more extensive data on, the normal behavior of amphibian populations. Populations tended to be thought of as defined by census units, which for amphibians are often breeding aggregations. A commonly held idea was that amphibians are highly philopatric as adults, returning to the same breeding site annually, and that this is often their natal site, making dispersal relatively uncommon (Sinsch 1990, 2006). Blaustein et al. (1994b, p 60) specifically mentioned that “due to the physiological constraints, relatively low mobility, and site fidelity of amphibians we suggest that many amphibian populations may be unable to recolonize areas after local extinction.” In parallel with work directed at the phenomenon of amphibian declines, more recent work on the movements and population dynamics of amphibians has indicated that many species may have much more complex spatial dynamics, occurring in metapopulations or even in unified populations using a variety of reproductive sites spread over a relatively large area (Sjogren-Gulve 1991, 1994; Skelly et al. 1999; Marsh 2001; Marsh and Trenham 2001; Trenham et al. 2001, 2003). By 1999, a substantial body of work had accumulated on the problem. A review by Alford and Richards (1999) cited 206 references, most published between 1980 and 1999, that addressed relevant topics. They separated nominated threats to amphibian diversity into 7 categories, and

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pointed out that most studies, while concentrating on single factors, acknowledged that interactions among factors were likely to be important. The complex nature of potential interactions has become increasingly evident in more recent literature, as will be discussed below and elsewhere in this volume. Alford and Richards also emphasized the need for more targeted studies at the population and metapopulation levels, and pointed out the difficulties inherent in interpreting the simple count data at breeding sites on which most amphibian monitoring had been based. The clear need for a better understanding of the global status of amphibians led the IUCN to organize and conduct a large-scale project, the Global Amphibian Assessment (GAA), to provide the best possible snapshot of the status of all amphibians (Stuart et al. 2004b). The world was separated into geographical regions, with a regional coordinator appointed to oversee the assessment process in each. Experts were involved in consultation processes in which every known species of amphibian was assessed against the IUCN criteria for species status (Stuart et al. 2004b). The results of the GAA indicated that a real crisis is occurring. The GAA assessed the status of 5711 species of amphibians. Almost one-third of the extant species (32.5%) were considered to be of conservation concern (IUCN categories “vulnerable” or higher; Stuart et al. 2004a). This is a substantially higher proportion than for either birds or mammals (Stuart et al. 2004a), although the status of 22.5% of species could not be assessed due to insufficient data. Stuart et al. (2004a) also pointed out that, unlike birds and mammals, the situation of amphibians appeared to be deteriorating rapidly, with the majority of known extinctions occurring late in the 20th century. Stuart et al. (2004a) attempted to examine trends by extrapolating the status of amphibians back to 1980. They used a conservative approach, incorporating any data that were available on population trends, changes in habitat, and conservation actions, with the default being that species had not changed since 1980 (Stuart et al. 2004b). Additional data accumulated when the GAA was partially updated in 2007 (IUCN 2008). A comparison of the overall numbers of species in each IUCN Red List category estimated for 1980 and observed in 2004 and 2007 appears in Table 2.1. The deterioration of the overall status of amphibians is most clearly illustrated by the dramatic increase in the numbers of species in the Table 2.1 Summary of the Status of Amphibians from the Global Amphibian Assessment (Stuart et al. 2004a, 2004b) and Its Recent Partial Update Number of Species

Percent Change

Category

1980

2004

2007

Extinct Extinct in the wild Total extant species Critically endangered Endangered Vulnerable (VU) Total VU and above Near threatened (NT) Least concern Data deficient

25 0 5718 231 807 734 2094 322 2322 1302

34 1 5709 427 761 668 2215 359 2199 1294

34 1 5881 441 737 630 2277 369 2277 1426

1980–2004 36.0 — –0.2 84.8 –5.7 –9.0 4.7 11.5 –5.3 –0.6

2004–2007 0.0 0.0 3.0 3.3 –3.2 –5.7 –2.6 2.8 3.5 10.2

Source: IUCN (2008). Note: The status of amphibians deteriorated markedly between 1980 and 2007. Although the numbers of species in some of the less threatened categories decreased, the table shows that this was due primarily to species moving upwards into the critically endangered and extinct categories. The small decrease in total NT and above in 2007 reflects a few species changing from vulnerable to lower threat categories based on improved information. The large increase in data-deficient species in 2007 largely reflects an increase in the rate of description of new species, particularly from tropical areas.

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“critically endangered” category in the interval between 1980 and 2004. Much of this occurred as species moved up from less threatened categories, causing the numbers in the categories of intermediate threat to decrease. The numbers in the “near threatened” (NT) category also increased, as species moved from “least concern” (LC) to NT. Even in the short interval from 2004 to the next partial update of the GAA in 2007, an additional 14 species were added to the “critically endangered” category. A few species were moved downward as additional data became available, and a substantial number of new species were described, mostly from poorly known areas in the tropics, causing an increase in the number of species regarded as “data deficient.” Stuart et al. (2004a) discussed the changes in status of amphibians between 1980 and 2004 in much greater detail, pointing out that a total of 435 species moved to higher categories of threat during this period. They also examined the sources of threats to these “rapidly declining” species, and found that the primary threat to 50 species was overexploitation, to 183 was habitat loss, and to 207 was “enigmatic declines,” which they defined as declining, even when suitable habitat remains, for reasons that are not fully understood. The distributions of these threats are geographically clustered, with overexploitation most common in the Asian region, habitat loss in North America, Europe, and Africa, and enigmatic declines predominant in Central and South America and Australia. The problem of amphibian declines is therefore clear; a substantially higher proportion of extant amphibians are vulnerable to extinction than are either of the 2 other terrestrial vertebrate groups for which a full assessment has been made. The best assessment to date suggested that approximately 8% of species had undergone rapid declines during the period 1980 to 2004. Many species are threatened by the clear anthropogenic factors of habitat loss and overexploitation, but welldocumented rapid declines to local extinction occurred during the 1980s and 1990s, at protected sites where habitat loss was not occurring — the “enigmatic declines” of Stuart et al. (2004a). Fortunately, during the period 2000 to present, a great deal of progress has been made toward understanding the causes of declines, including those labeled by Stuart et al. as enigmatic.

2.2 Potential Causes of the Amphibian Decline Crisis Most of the potential causes of amphibian declines were suggested early in the history of research into the problem. One of the earliest published summaries (Blaustein and Wake 1990) nominated habitat destruction, exotic predators, pollution, and utilization as food and pets as possible causes. Three more potential causes were quickly added to the list: effects of climate change (Herman and Scott 1992; Pounds and Crump 1994), disease, in combination with environmental factors (Carey 1993; Blaustein et al. 1994a), and increased levels of ultraviolet radiation (Blaustein et al. 1995). Disease acting alone, in the form of an invasive exotic pathogen, was suggested as a potential cause of widespread declines in eastern Australia by Laurance et al. (1996). Most recently, Harris et al. (2006) suggested that changes in the cutaneous bacterial assemblage of amphibians might increase their vulnerability to disease. From the beginning (Blaustein and Wake 1990), most authors acknowledged that causes were likely to be complex, involving interactions of more than a single factor. This was particularly true for climate change, which usually is suggested to act through changes in other factors. The process of linking specific threats to the species they affect is incomplete and often based largely on opinion. However, for most species under greater threat that have been more intensively investigated, there is now at least some understanding of the source of threats. Table 2.2 summarizes threats nominated for all species in the current update of the GAA (IUCN 2008). The proportion of species in each threat category subject to each nominated type of threat obviously differs across threat categories. This might be expected for the categories near threatened (NT), least concern (LC), and data deficient (DD), which have been studied to different degrees and for which nominated threats may reflect opinions or extrapolations. However, even when only the 3 most threatened categories, vunerable (VU), endangered (E), and critically endangered (CE) are considered, the source of threats differs significantly among categories (chi-squared contingency test, Χ2 = 121.4, 10 d.f., p < 0.0001). For easier visualization, these data are presented as percentages in Figure 2.1. This shows

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Table 2.2 Summary of Numbers of Species in Each IUCN Category Thought to Be Subject to Threats of Different Types Number of Affected Species in IUCN Category Threat

CE

E

VU

NT

LC

DD

Habitat loss Invasive species Utilization Pollution Disease Unknown threats Total number of species in category

145 62 30 168 200 3 441

308 91 43 213 125 7 737

211 68 48 197 79 18 630

130 55 35 110 32 25 369

457 149 104 401 54 867 2277

194 57 21 110 50 96 1426

Percent of species

Source: IUCN (2008). Note: Numbers within categories sum to more than the number of species in that category as some species are subject to multiple threats. CE = critically endangered, E = endangered, VU = vulnerable, NT = near threatened, LC = least concern, DD = data deficient.

50 45 40 35 30 25 20 15 10 5 0

disease pollution habitat loss invasive species utilization unknown threats

DD

LC

NT

VU

E

CE

IUCN category Figure 2.1  Percentage of amphibian species with extant populations in nature in each IUCN threat category for which 6 categories of threats are believed to be operating. Threat categories are arranged in order of increasing severity. From the left they are DD = data deficient, LC = least concern, NT = near threatened, VU = vulnerable, E = endangered, CE = critically endangered. Connecting lines are included to aid in visualizing how percentages change across categories. The total proportion for which at least one known threat has been nominated differs among categories, probably because greater effort has been focused on species under greater threat. The proportion of species known to be threatened by disease increases dramatically as the threat category increases. This is due mostly to the large number of species in the American tropics and Australia that have declined in association with outbreaks of the disease chytridiomycosis. (Data from IUCN 2008.)

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that the most common threat to species in the CE category is disease, threatening 200 species; most of these are threatened by the global emergence of the disease chytridiomycosis (Skerratt et al. 2007), which has repeatedly caused dramatic declines to local extinction of entire regional faunas (e.g., Lips et al. 2006). Figure 2.1 makes it clear that although this disease has led to dramatic declines and may be responsible for many of the transitions of species into the CE category that occurred between 1980 and 2007 (Table 2.1), other factors also represent major, ongoing threats to amphibian biodiversity. Recent work on each major threat category is discussed under separate headings.

2.2.1 Habitat Destruction The effects of habitat destruction are clear; when the major part of the natural habitat for a species is harvested or is converted to agricultural production or human habitation, species are negatively affected (Ash 1988; Semlitsch and Bodie 1998; Azevedo-Ramos and Galatti 2002). These effects can be felt even when remaining natural areas are set aside as reserves; populations in remaining fragments can be subject to extinction due to demographic stochasticity (Shaffer 1981; Lande 1993; Green 2000, 2003), and species living in metapopulations (Hanski 1994), or in integrated populations that shift their spatial distributions with changing habitat quality (Marsh and Trenham 2001; Bradford et al. 2003; Skelly et al. 2003; Smith and Green 2005), may be subject to extinction even when 20% or more of the original habitat is intact (Hanski 1994). More subtly, habitat destruction can have strong effects when it disrupts the ability of amphibians to undergo habitat shifts, as many species do either seasonally or at life history transitions (Becker et al. 2007). In the current GAA database, habitat destruction is the threat most commonly identified for species in all categories except CE (Table  2.2). However, clear examples of large-scale declines of amphibians caused by habitat destruction are rare in the literature. This may be because most broad-scale clearance of native habitat is presently occurring in regions where the amphibian fauna is relatively poorly known. For example, Crump (1971, 1974) documented an extraordinarily diverse amphibian fauna in the vicinity of Santa Cecilia, Ecuador, in an area that was subsequently heavily modified for agriculture (M. Crump, personal communication); most species disappeared locally from the modified habitat, but whether this substantially affected the total population of any species is not known. Habitat disturbance caused by mining in the Western Ghats of India reduced amphibian species richness by 50% (Krishnamurthy and Hussain 2004). Woinarski et al. (2006) showed that broad-scale clearance of native vegetation in Queensland, Australia, led to declines in many native vertebrates, including frogs, and an increase in the abundance of exotic invaders. Diniz et al. (2006) examined an area of the Brazilian Cerrado for possible compromises between preserving amphibian diversity and allowing development for human use, and found that although there is a positive correlation between the desirability of areas for human use and conservation, it might be possible to design a relatively effective set of reserves by conserving approximately 10% of the original habitat. When intensive habitat modification occurs in relatively small patches, amphibians that vacate those patches can often quickly recolonize them (Ash 1988, 1997; Alford and Richards 1999; Aubry 2000; Lehtinen and Galatowitsch 2001; Ash et al. 2003). On the other hand, even relatively minor changes in remnant patches in landscapes that are already heavily modified may tip the balance between persistence and extinction (Semlitsch and Bodie 1998; Babbitt et al. 2006). In addition to outright habitat destruction, amphibians can be affected by changes in habitat quality. Probably because amphibian populations have often been defined as the individuals using a particular breeding site (Marsh and Trenham 2001), many studies have focused on the quality of aquatic habitats and in some cases a narrow band of terrestrial habitat surrounding them. One of the simplest aspects of quality for the pool and pond habitats used by many amphibians is hydroperiod, which can strongly affect the composition of larval and breeding assemblages (Adams 1999; Eason and Fauth 2001; Baldwin et al. 2006; Seigel et al. 2006; Werner et al. 2007). Changes in hydroperiod have been suggested as causes of regional declines in some areas (Daszak et al. 2005; Palis et al. 2006). Changes in land use of surrounding terrestrial habitat can alter hydroperiods

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(Gray and Smith 2005), as can groundwater extraction (Guzy et al. 2006), and shifts in the hydroperiod of pools and ponds throughout entire regions due to global climate change may be a major threat to amphibians in the near future. These may be exacerbated by or interact with changes in the timing of seasonal reproduction caused by changes in temperature regimes (Blaustein et al. 2001; Chadwick et al. 2006). The quality of the aquatic habitat can also be modified biologically. Exotic animals and diseases will be considered separately, but 1 often overlooked factor is invasive aquatic vegetation. For example, Brown et al. (2006) showed that the invasive weed purple loosestrife (Lythrum salicaria) can negatively affect Bufo americanus tadpoles, probably through a combination of toxicity through ingestion of the leaves and alterations in the phytoplankton assemblage of pools infested with the weed. Many other aspects of water quality can affect the suitability of breeding habitats. Environmental contaminants can have drastic and interactive effects, as discussed more fully below and elsewhere in this volume. Other factors that are commonly found to affect the suitability of habitat include pH and aluminum (e.g., Anderson et al. 1999), pH conductivity, and depth (e.g., Babbitt et al. 2006), and combinations of these with multiple other factors (Brodman et al. 2003). In addition to water quality, the suitability of water bodies as reproductive sites is affected by the surrounding terrestrial habitat. These effects can act by altering the quality or quantity of water available, or changing the hydroperiod (Jansen and Healey 2003; Dayton and Fitzgerald 2006), but can also be caused by the suitability of the surrounding habitat for terrestrial juvenile and adult amphibians (Skelly 2001; Bradford et al. 2003; Knapp et al. 2003; Van Buskirk 2005; Baldwin et al. 2006; Otto et al. 2007). Many amphibians can occupy nonbreeding habitats that are not immediately adjacent to reproductive sites; when this occurs, connectivity between breeding sites and nonbreeding habitat is important (deMaynadier and Hunter 1999). Some disruption of habitat, for instance, by roads, can be tolerated by some species while adversely affecting others (deMaynadier and Hunter 2000). Becker et al. (2007) used the term “habitat split” to describe disruption of the connectivity between habitats used by differing life history stages, and showed a negative correlation between increasing habitat split and the species richness of amphibians with aquatic larvae in the Brazilian Atlantic Forest. They suggested that habitat split may be 1 reason why species with aquatic larvae are often suggested to be more subject to declines than are terrestrial breeders. Pond-breeding amphibian species occur in nested subsets at various spatial scales; those with greater specificity in terrestrial or aquatic habitat are more nested, as are poorer dispersers (Hecnar and M’Closkey 1997), a pattern suggesting that both selective extinction and selective colonization are responsible for patterns of occurrence at a regional scale. Habitat quality at multiple scales, within breeding sites and terrestrial sites, affects population density of European palmate newts (Triturus helveticus; Denoel and Lehmann 2006), and at least some amphibian species show distinct thresholds for habitat variables or combinations of variables at which density or occupancy changes abruptly (Denoel and Ficetola 2007). Homan et al. (2004) showed that threshold densities of forest cover vary depending on distance from breeding site for a frog (Rana sylvatica) and a salamander (Ambystoma maculatum); both species required substantial cover at relatively great distances from breeding sites (1 km). Houlahan and Findlay (2003) showed similar effects for a multispecies assemblage. Habitat loss and degradation are often continuous processes, so that slow loss of populations, and increasing exposure to exotics and invaders from modified habitats, can lead species to these thresholds (Hobbs and Mooney 1998). Cushman (2006) found that connectivity is an important determinant of the viability of regional amphibian faunas. He pointed out that the impacts of loss and fragmentation are likely to be felt first by more dispersive species, but that over longer timescales equal effects are likely on less dispersive species. Hazell (2003) suggested that conservation research in Australia should focus on how animals cope with modified habitats, and Johansson et al. (2005) demonstrated that effects can be complex and depend on factors that vary regionally. Gardner et al. (2007) pointed out that habitat change is associated with the greatest number of amphibian population declines, and that most knowledge of its effects comes from the relatively depauperate faunas and heavily modified habitats of North America and Europe. They suggested that more research needs to focus on its effects, particularly in areas of the world where habitat change is more rapid and extreme. Understanding the factors affecting the

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distribution and abundance of amphibians, and planning for their conservation, requires a detailed understanding of how they use habitat across a wide range of scales, from individual behavior, food, and microhabitat requirements (Altig et al. 2007) to regional vegetation structure and connectivity.

2.2.2 Exotic Predators Amphibians and their larvae are vulnerable to a wide array of predators (Hecnar and M’Closkey 1996; Alford 1999; Gunzburger and Travis 2005), many of which are likely to affect population dynamics and ultimately persistence (Kats and Ferrer 2003). Given the wide range of potential predators of amphibians, and the widespread introductions of predators that have occurred globally in both terrestrial and aquatic habitats (Kats and Ferrer 2003), the literature on possible effects of exotic predators on amphibian populations may be overly concentrated on the effects of fishes and exotic amphibians in aquatic systems. There appears to be very little information on possible effects of other aquatic predators, or on the effects of exotic predators on the terrestrial stages of amphibians (Ahola et al. 2006; Anthony et al. 2007). Predatory fishes have been widely introduced for sportfishing in many countries, and have strongly affected many of the systems they have occupied. They appear to be a major factor in declines of frogs in the California Sierra (Bradford 1989; Bradford et al. 1993; Fellers and Drost 1993; Drost and Fellers 1996; Knapp and Matthews 2000; Vredenburg 2004; Knapp et al. 2007). They are also implicated in declines of amphibians in other regions of North America (Monello and Wright 1999; Pilliod and Peterson 2001), and in southeastern Australia (Gillespie 2001; Hamer et al. 2002), Europe (Meyer et al. 1998; Martinez-Solano et al. 2003; Denoel et al. 2005; Orizaola and Brana 2006), and South America (Ortubay et al. 2006). The North American bullfrog, Rana catesbeiana, has been introduced to many areas outside its native range by the aquaculture industry. It is a voracious predator of larvae and often adults of other amphibians, and has been implicated in a number of declines (Hayes and Jennings 1986; Lawler et al. 1999; Monello and Wright 1999; Adams 2000; Doubledee et al. 2003; Bradford et al. 2004; Pearl et al. 2004). Other predators that may have caused declines include mink (Neovison vison; Ahola et al. 2006) and crayfish (Gamradt and Kats 1996; Cruz and Rebelo 2005; Cruz et al. 2006). In addition to their direct effects via the consumption of eggs, larvae, or adults, exotic predators can have a range of indirect effects. They may serve as vectors for parasites or pathogens (Daszak et al. 2004), and can increase the costs of antipredator behavior, reducing opportunities for growth and foraging (Chivers et al. 2001; Teplitsky et al. 2003; Cruz and Rebelo 2005). Their presence increases the frequency and intensity of stress responses, and may increase the effects of environmental contaminants (Relyea 2003, 2004, 2005; Teplitsky et al. 2005). Conversely, the effects of contaminants may alter antipredator behavior, making some species more vulnerable to predation (Bridges 1999). The effects of exotic predators may increase or decrease with broad-scale environmental changes associated with global climate change. Adams (1999, 2000) suggested that the spread of exotic predators (R. catesbeiana and sunfish) was facilitated by changes in hydroperiod toward greater permanence. Broomhall (2004) demonstrated that the thermal regime experienced by frog eggs can affect their sensitivity to environmental contaminants, and Rohr and Madison (2003) demonstrated that environmental moisture levels reduced the responses of newt efts to conspecific predator avoidance cues, suggesting that climate change might increase their vulnerability to predation.

2.2.3 Environmental Contamination Environmental contaminants are frequently suggested as potential primary factors or cofactors in amphibian declines (e.g., Alford and Richards 1999; Blaustein et al. 2003; Sih et al. 2004). There is strong correlational evidence from landscape-scale data that windborne pesticides have contributed to declines of Rana muscosa in the California Sierra Nevada (Davidson and Knapp 2007). Correlational evidence also suggests that declines, for example, those of Acris crepitans (Reeder

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et al. 2005) and Desmognathus fuscus (Bank et al. 2006), have been caused by surface-borne pollutants. The prevalence of such effects may well be underestimated; Schiesari et al. (2007) showed that tropical fauna, and species with limited geographic ranges, are understudied; if broad ranges are correlated with broad tolerances, this may mean that the literature is biased against both tropical species and species with narrower environmental tolerances. Effects of contaminants are discussed at length elsewhere in this volume, so the treatment here is brief, considering mostly studies that highlight the complex, interactive nature of the effects of environmental contamination. Many recent experimental studies have revealed that the effects of environmental contaminants are not well characterized by standard laboratory assays. This has 3 general causes. First, and simplest, the physical environment in the field differs from that in the laboratory; temperature, humidity, and the background chemistry of water do not mimic standard laboratory conditions, and any of these may alter the effects of contaminants. Second, many species are exposed to more than one, and often many, contaminants simultaneously, and their effects can interact in complex and unpredictable ways. Third, the fitness of animals in natural populations is determined by both their physical and their biological environments, and the effects of contaminants as mediated by the biological environment can be unpredictable and sometimes paradoxical. Temperature regime can determine how contaminants affect larval amphibians (e.g., Boone and Bridges 1999; Broomhall 2004). This implies both that more realistic toxicological studies should examine the effects of contaminants at a wider range of temperatures (Boone and Bridges 1999) and that the effects of contaminants may be altered by climate change (Broomhall 2004), which could lead to the emergence of impacts on species that have historically coexisted with contaminants. The effects of mixtures of contaminants can be highly unpredictable. For example, relatively low concentrations of carbaryl and nitrate, when either was applied alone, increased the rate of development and mass at metamorphosis of larval Rana clamitans (Boone et al. 2005). However, this positive effect disappeared when the contaminants occurred together. Hayes et al. (2006) examined 9 pesticides alone and in a variety of mixtures, and found that more complex mixtures generally had more negative impacts, both on life history characteristics such as growth rate and on the development of the endocrine system. The effects of environmental contaminants can interact with those of predation. Relyea (2003) demonstrated that the effects of carbaryl on larvae of several species increased, sometimes greatly, when they were also exposed to the stress caused by the presence of predatory newts. Predator cues also amplified the effects of the herbicide Roundup on larval Rana sylvatica (Relyea 2005). Conversely, exposure to the herbicide amitrole decreased the response of larval Bufo bufo to the chemical cues released as a predator fed on conspecifics (Mandrillon and Saglio 2007); this could increase the negative effects of predators on frog populations. A similar effect was suggested by Bridges (1999) and Teplitsky et al. (2005). Other interactions may reduce the effects of contaminants; for example, Relyea et al. (2005) showed that in mesocosms, the insecticide malathion, which can be toxic to tadpoles at high dosages, can increase their survival at low dosages by removing predatory insects. Boone and Semlitsch (2003) found a similar effect of carbaryl, which decreased the effect of predatory crayfish on larval Rana catesbeiana. However, higher concentrations of carbaryl negatively affected tadpole survival, outweighing the positive indirect effects of predator removal. Carbaryl had very similar effects in a study by Mills and Semlitsch (2004) on larval Rana sphenocephala. Davidson et al. (2007) convincingly demonstrated using correlational data that both windborne pesticides and introduced fishes have contributed to declines in Rana muscosa. As well as affecting the interactions of amphibians with predators, contaminants may also modify their interactions with diseases and parasites. This can occur via direct effects of contaminants on the amphibian immune system (Christin et al. 2003; Linzey et al. 2003; Houck and Sessions 2006). Interactions of contaminants with specific parasites and diseases have been documented in a variety of systems and are discussed in the following sections dealing with parasites and diseases. Contaminants can also interact with the effects of intraspecific and interspecific competition. Rohr et al. (2006) found that the immediate negative effects of exposure of larval Ambystoma barbouri to atrazine were partially ameliorated by reductions in intraspecific competition, although in the longer

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term, postexposure decreases in survival of exposed animals reduced this effect. Interspecific competition can also be reduced by contaminants; for example, carbaryl can decrease the competitive effects of zooplankton on tadpoles (Boone and Semlitsch 2001). Few experimental studies have set out to examine the effects of competition, predation, and contaminants simultaneously. Several studies by Boone et al. (Boone and Semlitsch 2001; Boone et al. 2004, 2007) have used experiments in mesocosms and larger systems to examine contaminant effects in complex assemblages. The outcomes of these experiments are complex and almost certainly specific to the systems examined, but they clearly support the general conclusion that the effects of environmental contaminants in natural systems are likely to be different in unpredictable ways from their effects in simple standardized assays.

2.2.4 Human Utilization The GAA database identifies utilization as a threat to between 5 and 10% of species in each threat category (Table 2.2, Figure 2.1). Most utilization is as food for human consumption or in the pet trade, although there is substantial use of amphibians for other purposes, such as in traditional medical preparations, in some regions. Frogs are commonly used as food in many parts of the world. There is a large international trade in fresh and frozen frogs and frogs’ legs. Many of the species used as food are large ranids (Dash and Mahanta 1993; Kusrini and Alford 2006; Tyler et al. 2007). Some of these species appear to be capable of sustained harvesting; Kusrini and Alford (2006) estimated that the annual frog harvest for human consumption in Indonesia is on the order of 100 million frogs, but suggested that the harvest currently appears to be operating within the bounds of sustainability. However, trade at a similar level (Pandian and Marian 1986) led India to ban the export of frogs’ legs in the 1980s as a factor threatening frogs and reducing their provision of ecosystem services in the form of insect control. The international trade in live frogs and frog parts for human consumption also poses a threat via the likely transport of pathogens (Berger et al. 1999; Daszak et al. 2003; Mazzoni et al. 2003; Rowley et al. 2007; Skerratt et al. 2007). Amphibians are also common in the pet trade. Documentation of the extent of this trade and its potential impact on amphibian populations is surprisingly difficult to come by. It is clearly massive, and much of it is so poorly documented that impacts may be impossible to assess (Schlaepfer et al. 2005). Schlaepfer et al. (2005) found that more than 2.5 million wild-caught amphibians and reptiles imported into the United States annually between 1998 and 2002 were not identified in official trade records to the species level. Harvesting for the pet trade was suggested by Andreone et al. (2005, 2006) to be a major threat to the endangered frogs of Madagascar. The harlequin frogs of the genus Atelopus, which have been heavily affected by epidemic outbreaks of chytridiomycosis, have also been heavily collected for export as pets (La Marca et al. 2005), as have many species of dendrobatids (Gorzula 1996). The pet trade may well pose an even greater risk for transporting exotic pathogens than the food trade does; although the number of animals transported is lower, the conditions in which the pet trade occurs probably make it likely that diseases and parasites can be transmitted to naïve hosts, and it is also more likely that pets may be accidentally or deliberately released into habitats to which they are exotic. Another form of utilization, with particularly negative indirect consequences, is the use of amphibians as live bait for fishing in some regions. This practice occurs in many areas of the United States, supplied by a large, and largely unregulated, interstate trade (Riley et al. 2003; Jancovich et al. 2005; Storfer et al. 2007).

2.2.5 Disease Some of the most rapid progress in understanding the causes of amphibian declines in the past decade has been made in the area of threats due to disease. Diseases and parasites are very unlikely

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to drive hosts with which they have long histories of coexistence to extinction (deCastro and Bolker 2005). However, the emergence of new infectious diseases, facilitated by rapid global transport and rapidly changing climatic conditions and ecological relationships, has recently been recognized as a major threat to the health of wildlife and humans (Daszak et al. 2001, 2003; Harvell et al. 2002). Emerging infectious diseases (EIDs) are defined as “diseases that are newly recognized, newly appeared in the population, or are rapidly increasing in incidence or geographic range” (Daszak et al. 2000, p 446). Even EIDs are likely to drive host species to local or global extinction only under certain conditions, including small host population sizes, the presence of pathogen reservoirs outside the affected host species, and possibly the specific behavior or social system of the affected host (deCastro and Bolker 2005). Disease is presently listed as a threat to 540 amphibian species (Table 2.2), and threatens the greatest proportion of species in the CE category (Figure 2.1). When amphibian declines are linked to EIDs, it is important to understand the causes of emergence (Rachowicz et al. 2005), since those are the ultimate causes of those declines. When diseases cause local extinctions, understanding how this can happen (deCastro and Bolker 2005) can suggest strategies for managing those diseases in other populations or species. Early suggestions relating amphibian declines to disease focused on known pathogens. Many local disease outbreaks recorded in North America and Europe prior to 1990 were attributed, often entirely on the basis of symptoms, to the disease red-leg. This can be caused by a variety of pathogenic bacteria, but was commonly attributed to Aeromonas hydrophila (Cunningham et al. 1996; Green et al. 2002). Because the symptoms of red-leg are very general, and are consistent with a variety of diseases, it seems likely that many of these outbreaks were caused by other diseases, such as ranaviral infections or chytridiomycosis (Green et al. 2002). Carey (1993) suggested that environmental stresses might compromise the immune function of amphibians, causing disease emergence due to increases in susceptibility to infections. The fungus Saprolegnia ferax, which attacks aquatic egg masses, experimentally decreased survival rates in frogs of the northwestern United States (Blaustein et al. 1994a; Kiesecker and Blaustein 1995), and has been suggested to possibly affect amphibians in other regions. Ranaviruses were isolated and identified from outbreaks in amphibian populations in several areas in the mid 1990s (Cunningham et al. 1996; Jancovich et al. 1997). Laurance et al. (1996) suggested that widespread outbreaks of an emerging pathogen might have been the proximate cause of local and global extinctions in a regional amphibian fauna. Although increased research on amphibian diseases has rapidly increased the number known (Carey 2000; Essbauer and Ahne 2001; Kiesecker et al. 2004), substantial research has been concentrated on 3 categories of disease that can produce mass mortality: ranaviruses and their relatives; saprolegniosis, caused by fungi in the genus Saprolegnia; and chytridiomycosis, caused by the amphibian chytrid fungus Batrachochytrium dendrobatidis. 2.2.5.1 Ranaviruses Ranaviruses appear to be widely endemic. Several ranaviruses have been described from wild and captive amphibians, including Ambystoma tigrinum virus (ATV; Jancovich et al. 1997) and Bohle iridovirus (BIV; Speare and Smith 1992). They are all closely related to fish viruses, for example, epizootic haematopoietic necrosis virus (EHNV) (Hengstberger et al. 1993; Yu et al. 1999; Cullen and Owens 2002). Some of the viruses in this group can be transmitted between amphibians and fishes (Ahne et al. 1997); however, at least ATV appears to be specific to salamanders (Jancovich et al. 2001). The species that can cross-infect both classes may have major potential for outbreaks, since there are many possible reservoirs and modes of transport (Ahne et al. 1997). Ranaviruses can cause large-scale die-offs in captive or laboratory populations (Cullen and Owens 2002). They have also caused die-offs in natural populations of salamanders (Jancovich et al. 2001; Brunner et al. 2004; Collins et al. 2004; Storfer et al. 2007) and frogs (Pearman and Garner 2005; Fox et al. 2006; Harp and Petranka 2006). They seem to cycle through local populations (Brunner et al. 2004), causing occasional epidemic outbreaks and mass mortality, but do not appear to be major threats to the persistence of any known amphibian species. This may change in

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the presence of environmental contaminants; Forson and Storfer (2006) demonstrated that larval Ambystoma tigrinum that were exposed to atrazine were more susceptible to ATV infection than unexposed larvae, suggesting that contaminants might alter the dynamics of the disease to favor outbreaks. There is also a danger of larger-scale outbreaks caused by transport of animals or contaminated water. Harp and Petranka (2006) showed that placing sediments contaminated with ranavirus in mesocosms transmitted the infection to tadpoles. This suggests that uncontrolled disposal of water used to transport amphibians or tropical fishes may pose a serious risk of long-distance transport and introduction of exotic ranaviruses, which could have serious effects on naïve species or populations of amphibians. 2.2.5.2  Saprolegnia and Other Fungal Pathogens Fungi of the genus Saprolegnia have long been known to attack fishes (Blaustein et al. 1994a). They can attack larval amphibians (Bragg and Bragg 1958; Walls and Jaeger 1987), but the greatest concern has been raised over their effects on eggs. Blaustein et al. (1994a) documented an outbreak of the fungus Saprolegnia ferax that killed an estimated 95% of the eggs deposited by a large population of Bufo boreas in Oregon. Kiesecker and Blaustein (1995) reported strong effects of the fungus on eggs of both B. boreas and Rana cascadae; these appeared to be stronger in eggs exposed to higher levels of UV-B radiation. Some negative effects of the fungus may be compensated by decreased density dependence in the larval stage; Kiesecker and Blaustein (1999) found that in artificial pond experiments, Rana cascadae that were exposed to S. ferax as eggs but survived to hatching grew and developed faster than individuals that were not so exposed, and therefore experienced higher larval densities. Kiesecker et al. (2001) suggested that Saprolegnia might have an increasing effect on amphibians as decreased water depths at oviposition sites caused increased exposure of eggs to UV-B radiation. The effects of Saprolegnia may be modified by predation on the fungus; Gomez-Mestre et al. (2007) found that larval Rana sylvatica grazed on and removed fungus from the eggs of Bufo americanus, increasing their survival. Other fungi, including Aphanomyces sp., also infect tadpoles (Berger et al. 2001). The fungus Mucor amphibiorum has produced epidemic outbreaks in captive populations of adult frogs (Creeper et al. 1998); however, only isolated cases have been observed in the wild (Speare et al. 1994; Berger et al. 1997). 2.2.5.3  Amphibian Chytrid Fungus Berger et al. (1998) showed that chytridiomycosis was the proximate cause of population crashes in Australia and Central America. The pathogen was described by Longcore et al. (1999) as Batrachocytrium dendrobatidis. It has become widely accepted (Rachowicz et al. 2005) that chytridiomycosis is the proximate cause of many of the amphibian declines categorized by Stuart et al. (2004a) as “enigmatic.” 2.2.5.3.1  Characteristics of the Pathogen Chytridiomycosis has caused population declines and extinctions across most of an entire class of vertebrates (Daszak et al. 1999). The very broad host range of B. dendrobatidis, including larvae and adults of many species that are resistant to chytridiomycosis as well as those of many species that are vulnerable to the disease, may explain its ability to drive populations of many species to extinction, since it can persist in less affected reservoir hosts (Hanselmann et al. 2004; deCastro and Bolker 2005; Woodhams and Alford 2005; Smith et al. 2007). The thalli of B. dendrobatidis live in the epidermis, which becomes infected by contact with waterborne flagellated zoospores. Propagation occurs by producing zoosporangia that release zoospores to the external environment. The development of chytridiomycosis thus requires multiple generations of successful colonization and recolonization of the host by propagules (Carey et al. 2006). In vitro, the fungus usually exhibits endogenous development (James et al. 2000), with zoospores encysting on a substrate, developing a network of rhizoids, and subsequently developing

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a zoosporangium, or sometimes several zoosporangia. On living amphibians, the zoosporangia begin developing within cells in the deeper layers of the epidermis and complete their development as those cells migrate toward the skin surface. The details of the process of encystment and initial infection on living amphibians have not been documented. However, to reach the interior of cells deep in the epidermis, it appears that B. dendrobatidis must switch to an alternate developmental pathway, exogenous development (James et al. 2000), in which the zoospore encysts on the host surface, then develops a germ tube through which the nucleus migrates into the interior of a host cell or tissue (James et al. 2000). In culture, encysted zoospores can develop apparent germ tubes extending toward isolated pieces of frog skin (J. Longcore, personal communication). The ability to facultatively switch developmental mode appears to be unusual for fungi of the phylum Chytridiomycota (Powell and Koch 1977). The flexibility exhibited by B. dendrobatidis may suggest that it regularly exploits a variety of substrates in nature. Johnson and Speare (2005) showed that B. dendrobatidis can survive for extended periods in the laboratory on a variety of potential environmental substrates, and Lips et al. (2006) found B. dendrobatidis DNA on 1 of 9 haphazardly chosen stream boulders. More work is needed on alternative substrates, to increase understanding of how B. dendrobatidis persists in the environment and how it may be dispersed. This may be aided by techniques recently developed for sampling environmental water for the presence of B. dendrobatidis DNA (Kirshtein et al. 2007; Walker et al. 2007). 2.2.5.3.2  Emergence and Persistence of Chytridiomycosis At present, B. dendrobatidis is widely distributed in Africa (Weldon et al. 2004; Goldberg et al. 2007), Europe (Mutschmann et al. 2000; Bosch and Martinez-Solano 2006), North America (Green and Muths 2005; Ouellet et al. 2005; Longcore et al. 2007; Pearl et al. 2007), Central and South America (Ron 2005; Carnaval et al. 2006; Lips et al. 2006; Puschendorf et al. 2006b), and Australia (Drew et al. 2006). It also occurs on many islands, both in the Atlantic and Carribbean and in the Pacific (Bell et al. 2004; Burrowes et al. 2004; Beard and O’Neill 2005; Diaz et al. 2007). It has not yet been detected in continental Asia; however, very few surveys have been done there (Rowley et al. 2007). Debate regarding the reasons for the emergence of chytridiomycosis has centered on 2 competing hypotheses, which Rachowicz et al. (2005) described as the novel and endemic pathogen hypotheses. The novel pathogen hypothesis postulates that B. dendrobatidis has recently greatly increased its geographic range, encountering naïve hosts that have no defenses against it. The endemic pathogen hypothesis in its simplest form suggests that B. dendrobatidis has historically had a wide geographic distribution, and has emerged as a major pathogen of amphibians due to widespread ecological changes, possibly associated with climate change, that have “tipped the balance” in the host-pathogen relationship (e.g., the climate-linked epidemic hypothesis of Pounds et al. [2006]). Various authors have suggested that the true causes may be a combination of local or global range expansion with environmental changes that favor the pathogen (Rachowicz et al. 2005; Blaustein and Dobson 2006). Genetic studies might provide a definitive answer to the question of the pathogen’s origin. However, to date, they have not been conclusive. Morehouse et al. (2003) used multilocus sequence typing to examine levels of nuclear genetic diversity among 35 strains of B. dendrobatidis from North America, Africa, and Australia. They found very low levels of variation, nearly constant frequencies of heterozygous genotypes, and little geographic patterning, suggesting that most reproduction is clonal and that the pathogen has recently dispersed globally. Morgan et al. (2007), however, examined 15 variable nuclear loci and found evidence for both recent dispersal and endemism in the Sierra Nevada of California. Berger et al. (2005a) showed that the virulence of B. dendrobatidis differed significantly among 3 strains, and Piotrowski et al. (2004) demonstrated differences in growth rate and sensitivity to pH among strains grown in vitro. These strain-specific differences might reflect genetic differentiation among strains.

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Two studies have examined large samples of museum specimens to look at the historical occurrence of B. dendrobatidis. Ouellet et al. (2005) examined specimens from 25 countries collected between 1895 and 2001, the great majority of which were collected after 1960 from North America. The earliest specimens they found infected with B. dendrobatidis were Rana clamitans collected in Quebec in 1961. They found that the prevalence of B. dendrobatidis infections in Quebec had not changed significantly between the decades 1960 to 1969 and 1990 to 1999, and suggested that the organism had been widespread in North America since at least the early 1960s. They noted that the number of specimens collected prior to 1960 that they had examined was not sufficient to establish with any certainty that the pathogen was absent before that time. Weldon et al. (2004) examined 697 specimens of Xenopus spp. collected in southern Africa between 1879 and 1999. The overall prevalence of B. dendrobatidis in the sample was 2.7%; there was no statistically significant trend in prevalence over time. They suggested that this established that the fungus is endemic to South Africa, in a stable association with Xenopus species, and that the disease might have been spread worldwide by the extensive trade in Xenopus. This untested hypothesis has frequently been cited as an established fact in the public media, and may be contradicted by the lack of genetic variation among African isolates (Morehouse et al. 2003) and the recent occurrence of epidemic outbreaks of chytridiomycosis in southern Africa (Hopkins and Channing 2003). Far more work needs to be done before any hypothesis is accepted; for example, a strong case could also be made for the hypothesis that the fungus originated in North America. It is widespread there, occurs in apparently stable associations with nondeclining species (Daszak et al. 2005), has been found in relatively early museum records, and shows some genetic differentiation on a local scale (Morgan et al. 2007). It could have been disseminated globally via the trade in bullfrogs (Garner et al. 2006). Laurance et al. (1996) examined the spatiotemporal pattern of declines in Queensland, Australia, and suggested that an epidemic disease was traveling northward at approximately 100 km·year–1. Their analysis was criticized by Alford and Richards (1997), who pointed out that their conclusions were based on 3, or at best 4, clusters of decline sites that could not be regarded as independent within clusters, and that within the largest cluster, spanning approximately 380 km, there was no evidence of a wavelike progression of outbreaks. The wavelike nature of outbreaks has also been emphasized in a series of publications by Lips (Lips 1999; Lips et al. 2006, 2008), in which the timing of outbreaks of chytridiomycosis leading to declines of montane amphibians in Central and South America was suggested as most consistent with a hypothesis of at least 4 separate introductions of B. dendrobatidis in Central and South America, followed by spread of the pathogen across the landscape. The pathogen presently occurs in almost all suitable habitats in Australia, Europe, and North, Central, and South America. Many of the sites at which it now occurs are remote, and have no introduced amphibian populations within at least hundreds of kilometers. If it originated in southern Africa (Weldon et al. 2004), B. dendrobatidis is capable of extremely rapid and efficient dispersal through a wide variety of relatively undisturbed habitats. This makes it surprising that it remained confined to certain frog species in southern Africa for its entire preceding evolutionary history. The endemic pathogen hypothesis also is supported by some lines of reasoning and evidence. Epidemic outbreaks do not always occur within weeks or months of the organism first appearing at a site. DiRosa et al. (2007) reported that B. dendrobatidis was present at a site in Italy for several years before an epidemic outbreak occurred. Richards et al. (1993) documented declines at Kirrama, Queensland, in the highly susceptible species Litoria rheocola approximately 1 year after the first known date of occurrence of B. dendrobatidis in that area (Berger et al. 1999). Puschendorf et al. (2006a) showed that B. dendrobatidis was widely distributed in Costa Rica as early as 1986, at least a year before the first known declines associated with chytridiomycosis in Costa Rica (Pounds and Crump 1994). Despite the widespread occurrence of B. dendrobatidis in many species in North America (e.g., Ouellet et al. 2005; Longcore et al. 2007; Pearl et al. 2007), only localized epidemic outbreaks have been reported, and amphibian populations are known to have coexisted with the pathogen for many years (Daszak et al. 2005).

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In regions where the pathogen is now endemic, the environment strongly affects its prevalence and the probability of epidemic outbreaks. In northeastern Queensland, Australia, for example, where all known populations above approximately 400 m elevation of 7 susceptible species were extirpated by epidemic outbreaks of chytridiomycosis during the late 1980s and early 1990s, no populations below that elevational cutoff were affected (McDonald and Alford 1999). The prevalence of B. dendrobatidis in Queensland frogs now fluctuates seasonally in a manner consistent with those elevational effects (McDonald et al. 2005; Woodhams and Alford 2005); it is higher in the cool, dry winter months and lower in the warm, wet summer months. Similar seasonal and elevational effects have been reported in other areas (Ron 2005; Longcore et al. 2007). Environmental temperature is probably 1 cause of these patterns. Piotrowski et al. (2004) examined the growth rates of B. dendrobatidis populations in vitro and showed that the pathogen grew between 4 and 25 °C and reproduced most rapidly between 17 and 25 °C. Temperatures above 30 °C killed the fungus. Woodhams et al. (2008) found that although development of the fungus slowed at temperatures below 17 °C, the numbers of zoospores produced by each thallus increased, so that rapid population growth on the host can probably be maintained over a wide range of temperatures between approximately 10 and 25 °C. They also found that sudden temperature decreases stimulated the release of zoospores; in nature, synchronous release of zoospores by many zoosporangia on many infected individuals could result in large peaks in the rate of acquisition of new infections during the cooler months. Several studies have produced correlational evidence that the timing of epidemic outbreaks of chytridiomycosis is controlled by weather, which may be changing with the global climate. Pounds et al. (2006) demonstrated a relatively strong relationship between increased air temperatures and the last year observed for species of Atelopus. They suggested that the apparent effects of higher temperatures might be caused by increases in cloud cover, which retains heat at night, causing amphibians to experience temperatures within the thermal range optimal for growth of B. dendrobatidis over longer periods of time. Bosch et al. (2007) showed a correlation between increasing temperatures and outbreaks of chytridiomycosis in Spain, and Laurance (2008) found that frog declines in upland Australian rainforests tended to occur during warmer periods. Outbreaks of chytridiomycosis in Atelopus species were linked by Lampo et al. (2006) to a severe dry season. Studies at the individual level provide some insight into the possible mechanisms of the effects of weather and climate. Woodhams et al. (2003) found that 16 hours of exposure to temperatures of 37 °C cured all infected Litoria chloris, while 16 hours at 8 °C caused infections to progress more slowly than they did in frogs housed at a constant 20 °C. Rowley (Rowley 2006; Rowley and Alford 2007b) tracked frogs of several sympatric species in the Australian Wet Tropics, and found they used the environment in ways that exposed them to very different moisture and temperature microenvironments. Both across species and within species, individuals that attained body temperatures above 25 °C had lower probabilities of infection by B. dendrobatidis. Rowley and Alford (2007a) demonstrated that the behavior patterns of tracked frogs were also likely to affect the probability of transmission of B. dendrobatidis. Taken together, individual level studies suggest that in the field, infections on individuals of many species may be maintained at relatively low prevalence and intensity by various combinations of dry environmental conditions, which inhibit the release of zoospores and thus reduce the rate of reproduction of the fungus; and higher body temperatures, which slow the growth rate of the fungus, may increase the effectiveness of immune responses, and may clear infections at temperatures above 30 °C. If this is the case, periods of overcast weather and high humidity may increase the growth rates of B. dendrobatidis populations on infected individuals, releasing large numbers of infective zoospores into the environment and producing epidemic outbreaks. Whatever the history of dispersal of B. dendrobatidis has been, it may presently be absent from some areas (Rowley et al. 2007), and it is clear that strong precautions should be taken to avoid aiding its dispersal by anthropogenic means (Skerratt et al. 2007).

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2.2.5.3.3  Determinants of Vulnerability to Chytridiomycosis In epidemic outbreaks and when it is present as an endemic, vulnerability to chytridiomycosis caused by B. dendrobatidis varies widely among species. Initial epidemic outbreaks often drive populations of some species to local extinction, cause others to decline but not to extinction, and leave others apparently unaffected (Richards et al. 1993; Lips 1999; Fellers et al. 2001; Bell et al. 2004; Bosch and Martinez-Solano 2006; Lips et al. 2006). Several studies have looked for correlations between the ecological and life history characters of amphibians and vulnerability to declines caused by chytridiomycosis. At a national Australian workshop in late 1997, McDonald and Alford (1999) presented an analysis of patterns of frog declines in eastern Australia showing that species more tightly associated with streams were more likely to suffer declines. This pattern has repeatedly emerged in subsequent analyses of tropical rainforest amphibians (Williams and Hero 1998; Lips et al. 2003; Hero et al. 2005). Several studies have also found correlations between the severity of the effects of chytridiomycosis and other characteristics, including geographic range size, body size, and fecundity (Williams and Hero 1998; Lips et al. 2003; Hero et al. 2005). These relationships are difficult to interpret, since when local extinctions associated with chytridiomycosis have been documented, they typically involve rapid and complete mortality of all terrestrial members of populations (McDonald and Alford 1999; Lips et al. 2006), rather than the slow spiraling toward extinction that might be expected from low fecundity and limited geographic range. Hamer and Mahony (2007) pointed out that Litoria aurea, which has suffered severe declines in association with chytridiomycosis, has life history characteristics more usually associated with invasive species than declining ones. Only Murray and Hose (2005) have examined correlates of decline in frogs using phylogenetically independent contrasts. They found that only geographic range size was correlated with the probability of decline across a large number of clades. Immune function can affect vulnerability to B. dendrobatidis and other diseases. Carey (2000) suggested that disruption of amphibian immune function by environmental stressors might be a factor in the emergence of disease. It appears that the adaptive and cellular immune systems of amphibians do not show strong responses to B. dendrobatidis, even in advanced stages of chytridiomycosis (Pessier et al. 1999; Berger et al. 2005b). Amphibians also possess innate immune defenses, in the form of antimicrobial peptides (AMPs) that are secreted by the granular glands onto the skin surface. Rollins-Smith et al. (2002a, 2002b) demonstrated the effectiveness of a variety of AMPs against B. dendrobatidis. Further studies (Rollins-Smith et al. 2003, 2005; Rollins-Smith and Conlon 2005) showed that many amphibian species produce one or more AMPs that can strongly inhibit the fungus. Woodhams and his co-workers (2006a, 2006b, 2007a) have produced strong correlational evidence that vulnerability of amphibian species to population declines caused by B. dendrobatidis is related to the effectiveness of their AMPs against the fungus. Davidson et al. (2007) showed that levels of AMP production in Rana boylii were decreased by exposure to the insecticide carbaryl. Much more work is needed, for example, to understand whether AMP production responds to infection or is purely constitutive, whether the chemical composition of AMP secretions varies among individuals and populations and whether it is affected by environmental conditions, and how this relates to susceptibility to chytridiomycosis. Another factor that can affect vulnerability to chytridiomycosis is interactions between B. dendrobatidis and other microbes that inhabit amphibian skin. The skin of many amphibians supports a complex microbial assemblage, with which the zoospores of B. dendrobatidis must interact during the infection process (Belden and Harris 2007; Culp et al. 2007). Changes in the composition of this assemblage are likely to alter these interactions; some assemblages may exclude the pathogen while others may be readily invaded by it (Belden and Harris 2007). Harris et al. (2006) demonstrated that several genera of bacteria commonly isolated from salamanders of 2 species inhibited the growth of B. dendrobatidis in culture. The composition of the bacterial assemblages on Plethodon cinereus and Hemidactylium scutatum and the nature of their interactions with fungi were explored in greater detail by Lauer et al. (2007, 2008). One of the bacterial metabolites that can inhibit B. dendrobatidis

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was identified and characterized by Brucker et al. (2008). Two other bacterial metabolites occur on the skin of P. cinereus at concentrations sufficient to completely inhibit the growth of B. dendrobatidis in vitro (R. N. Harris, personal communication). Cutaneous bacteria can also contribute to the defenses of frogs against B. dendrobatidis; Woodhams et al. (2007c) found that a significantly greater proportion of individuals of the threatened species Rana muscosa carried bacteria with activity against B. dendrobatidis in a population that had coexisted with the pathogen for 6 years than in a population that was declining due to chytridiomycosis. Harris et al. (2009) experimentally demonstrated that Plethodon cinereus that had been exposed to the bacterium Pseudomonas reactans suffered fewer negative effects after exposure to B. dendrobatidis than salamanders that had not been exposed to the bacterium. It is likely that the AMPs of amphibians interact with and modify their skin microbiota (Woodhams et al. 2007b). Understanding the complex ecology of the microbiota of amphibian skin is at a very early stage, but may lead to an ability to manipulate interactions, for example, through probiotic applications or introductions of bacteria that tip the balance against B. dendrobatidis (Woodhams et al. 2007b; Harris et al. 2009).

2.2.6 Amphibian Limb Malformations and Their Relation to Parasitic Disease Reports of possible increases in the incidence of developmental abnormalities, particularly limb malformations, in amphibians have attracted wide popular attention (Johnson et al. 2001, 2003; Loeffler et al. 2001; Chapter 16, this volume). Several factors can cause limb abnormalities in amphibians (Loeffler et al. 2001). These include environmental contaminants (La Clair et al. 1998; Natale et al. 2000; Qin et al. 2005; Taylor et al. 2005; Papis et al. 2006; Piha et al. 2006; Webb and Crain 2006), injuries to developing limb buds (Loeffler et al. 2001; Johnson et al. 2006), the direct effects of UV-B radiation (Starnes et al. 2000), and encystment of parasitic trematodes (Johnson et al. 1999, 2002; Kiesecker 2002). Johnson et al. (1999) demonstrated that cercariae of the trematode Ribeirioa sp. (later assigned to Ribeiroia ondatrae; Johnson et al. 2004) attacked larvae of the frog Hyla regilla, encysting near the developing hindlimb buds and producing hindlimb malformations similar to those found at field sites. Several subsequent studies by Johnson et al. (2001, 2002, 2003) demonstrated that Ribeiroia infections are a common, though not universal, cause of limb malformation in amphibians, and that it is likely that the prevalence of Ribeiroia-induced malformations increased in the late 20th century. Johnson et al. (2003) suggested that amphibian limb deformities caused by parasitic infection can be regarded as an emerging infectious disease. Johnson and Chase (2004) suggested that the emergence of infection by Ribeiroia spp. might be due to changes to aquatic food webs, in which nutrient runoff causes eutrophication, which causes a shift in the composition of the snail assemblage in ponds toward the genus Planorbella, which are intermediate hosts of Ribeiroia, leading to higher levels of infective cercariae in the water column and higher rates of limb malformations in frogs emerging from eutrophic water bodies. This hypothesis was experimentally demonstrated to be feasible by Johnson et al. (2007), who also showed that infection by Ribeiroia decreased survival of larval amphibians, as well as increasing rates of limb malformation. In addition to the effects of eutrophication, environmental contaminants present in agricultural runoff may affect the prevalence and intensity of trematode infections in larval amphibians, possibly by affecting immunocompetency (Kiesecker 2002). Although infection by Ribeiroia spp. is clearly not the only cause of amphibian malformations (Skelly et al. 2007), it appears to be a common one, with a pattern of emergence that is strongly linked to anthropogenic effects on the environment via eutrophication.

2.2.7 Ultraviolet Radiation One of the early and obvious signs of global-scale human influences on the climate and environment was increases in levels of incident ultraviolet radiation caused by changes in the outer layers

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of the atmosphere (Kerr and McElroy 1993). Ultraviolet-B (UV-B) radiation can directly reduce the survival of eggs and embryos of amphibians that rely on egg masses that remain near the surface of shallow water bodies (Ovaska et al. 1997; Lizana and Pedraza 1998; Broomhall et al. 2000; Hakkinen et al. 2001). In many systems, current evidence suggests that ambient UV-B levels are not high enough to cause such damage (Langhelle et al. 1999; Merila et al. 2000; Starnes et al. 2000; Calfee et al. 2006). However, evaluating direct effects on egg and embryo survival may not be sufficient; Pahkala et al. (2001) found that effects of higher levels of UV-B exposure only appeared in the larval stage, where they found a higher rate of developmental abnormalities and slower rates of growth and development in Rana temporaria that had been exposed to enhanced UV-B as embryos. Belden and Blaustein (2002) found similar effects on Rana aurora exposed to full ambient levels of UV-B in the field. A great deal of protection from environmental UV-B can be provided by the low transparency of many natural water bodies at these wavelengths (Palen et al. 2002) and by the optical properties of the jelly capsules of many amphibian eggs (Licht 2003). The direct effects of UV-B on eggs and embryos are likely to be mediated by the activity of repair enzymes such as photolyase (Blaustein et al. 1996, 1999; van de Mortel et al. 1998), which can change in response to changes in UV-B exposure (Smith et al. 2000). Experimental work on Patagonian frogs, which are exposed to increased levels of UV-B radiation due to ozone depletion (Perotti and Diegeuz 2006), suggested that melanin levels in eggs and embryos increase in response to increased UV-B exposure, but these increases may be insufficient to eliminate increases in malformation rates when UV-B exposure is high. In addition to its effects on aquatic stages of amphibians, UV-B could directly or indirectly affect terrestrial individuals. High levels of UV-B may damage the immune system (Carey et al. 1999), and might increase the susceptibility of individuals to disease (Kiesecker et al. 2001). Kiesecker et al. (2001) showed, using correlative and experimental data, that increases in exposure to UV-B increase the vulnerability of frog embryos to the pathogenic fungus Saprolegnia ferax. However, Garcia et al. (2006) found no evidence for any interaction between the effects of UV-B and chytridiomycosis on 3 species of western North American frogs. Evidence is also accumulating that damage caused by UV-B may interact, sometimes multiplicatively, with the effects of environmental toxins and other threats (Blaustein et al. 2003). The effects of high levels of nitrate fertilizers on the growth and development of larval amphibians can depend in a complex way on levels of exposure to ambient UV-B (Hatch and Blaustein 2003). Fite et al. (1998) showed that high ambient levels of UV-B can cause retinal damage in adult frogs, and Blaustein et al. (2000) demonstrated that exposure to UV-B changed the overall activity levels of newts (Taricha granulosa). Increasing levels of UV-B exposure decreased the expression of antipredator behavior by juvenile toads (Bufo boreas) and frog tadpoles (Rana cascadae, Kats et al. 2000); in nature, this should lead to decreases in survival and recruitment. Some species may be capable of responding evolutionarily to increases in ambient UV-B; Weyrauch and Grubb (2006) found that populations of Rana sylvatica with lower genetic diversity experienced higher larval mortality rates when exposed to UV-B than did populations with higher levels of genetic diversity. Although it appears that present levels of ambient UV-B are not a major cause of amphibian declines, it may have a role in some systems. That UV-B exposure can interact in unpredictable ways with other threats indicates that a much fuller understanding of its effects would be useful in the future management of threatened amphibian populations.

2.2.8 Climate Change Most of any effects of global climate change on amphibians that may have occurred to date are likely to have arisen through interactions of changed environmental conditions with other factors, many of which are discussed above. Most herpetologists appear to agree with Carey and Alexander (2003), who pointed out that it is unlikely that the simple thermal or other effects of recent climate change have been sufficient to explain recent declines in amphibian populations. However, effects of

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climate change can certainly cause changes in the abundance of amphibians in particular localities or habitats. Whitfield et al. (2007), for example, documented a long-term decrease in the abundance of amphibians at LaSelva, Costa Rica, and linked this change to climate-driven decreases in the quantity of leaf litter. There is evidence (Seimon et al. 2007) that some species of amphibians have extended their ranges to higher elevations due to deglaciation. Climate change within the ranges predicted for the near future may have severe impacts on many amphibian populations without the need for interactions with other factors. Williams et al. (2003), for example, suggested that an increase of 3.5 °C in average temperatures could cause the extinction of most of the vertebrates endemic to the Australian Wet Tropics, including many amphibians. There may be statistical problems with some of these predictions (Pearson and Dawson 2003; Dormann 2007). In addition, they assume that species’ ranges are presently limited only by environmental factors, and that the nature of these limits will not change as the environment changes. They thus ignore the well-established effects of history, evolution, and biotic factors (e.g., Williams and Pearson 1997; Graham et al. 2006; Dormann 2007; Heikkinen et al. 2007) on species’ ranges. It is also clear (Hauselberger and Alford 2005; Rowley and Alford 2007a) that sympatric amphibian species can experience very different environmental regimes due to microhabitat selection and other behavioral differences. Taken together, these suggest that accurate predictions of responses to climate change are not possible at present. However, it is very clear that if the environment changes sufficiently, there will ultimately be large effects on amphibians and on all other elements of the planet’s biota. More detailed knowledge of how amphibians experience and interact with the physical environment in the field should aid in refining the predictions of climate-based models, and in focusing conservation efforts to minimize as much as possible the effects of climate change on amphibians.

2.3 Management of Amphibian Populations Managing and conserving amphibian populations has been greatly hindered by the relative scarcity of information on their ecology at the individual and population levels. Early regional and national plans (e.g., Tyler 1997) included large components of research; proposed actions focused on habitat conservation and, in some cases, preliminary work toward the ability to carry out management in captivity. Substantial progress has been made toward a fuller understanding of many of the threats to amphibians, including environmental contaminants, UV-B, disease, and habitat modification, since these early plans were adopted, and it is likely that many of them need to be redrafted in the light of current knowledge. The Global Amphibian Conservation Action Plan (Gascon et al. 2007) may serve as a useful template, since it incorporates very current input from many conservation specialists and researchers. The “Summary of Action Steps” on pages 6 to 11 of that document presents most of the considerations necessary for a well-thought-out conservation plan. The remainder of the action plan provides a very useful review of the conservation biology of amphibians. At present, when species are imminently threatened, either by unknown causes or by outbreaks of chytridiomycosis, management options are limited. Mendelson et al. (2006) argued strongly in favor of short-term captive management as a tool to “buy time” for critically threatened species, and this idea has been adopted for some species. There is still a need for a much greater understanding of the basic behavior and ecology of many species (Biek et al. 2002), particularly those in the more poorly studied regions and faunas (Schiesari et al. 2007). The rapidly growing field of microbial interactions on amphibian skin offers some hope of eventually managing populations threatened by the amphibian chytrid fungus in the field (Woodhams et al. 2007b). It is very clear that habitat conservation must remain an important part of efforts to conserve amphibians (Gardner et al. 2007), and that it is essential to control human-assisted movements of organisms of all sorts (Kats and Ferrer 2003; Skerratt et al. 2007; Tyler et al. 2007). Failing to do that is difficult to understand, because unlike many other problems, solutions exist and are technically achievable. It is also crucial to continue to document the diversity of amphibians. Parra et al. (2007) point out that the number of known amphibian species has increased by approximately 40% since 1987, and

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that as many as half of the world’s species may currently be undescribed. This is a far higher rate of discovery and proportion of undescribed taxa than occurs in any of the other groups of terrestrial vertebrates. The process of status assessment also needs to be streamlined and made ongoing, rather than episodic (Stuart 2007).

2.4 Summary and Conclusions In less than 20 years, the phenomenon of global amphibian declines has moved from initial recognition to be one of the better-studied aspects of what is clearly a global crisis for biological diversity of all sorts (Gascon et al. 2007). Amphibians appear to be the most threatened class of vertebrates (Stuart et al. 2004a). The picture is not entirely gloomy, however. Research carried out since the problem was first identified has enormously deepened our understanding of the biology of amphibians and the nature of the factors threatening them. Two or more generations of planning and responses have begun to converge on action plans that should be possible to implement and should aid greatly in the conservation of many species. Entire fields of research that were largely or entirely unknown in 1990 have opened up and begun to be applied to solving the problem of amphibian declines. Much more study is still needed, but as the public has become aware of the extent of the problem, and the challenges it poses, the resources available to apply to the problem have begun to increase. The total budgets suggested in all chapters of the global amphibian conservation action plan (Gascon et al. 2007) amount to less than the costs of 2 large commercial airliners. It is to be hoped that the global community will perceive that preserving several thousand species of amphibians is worth that much.

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The Global Status of Reptiles and Causes of Their Decline Brian D. Todd, John D. Willson, and J. Whitfield Gibbons

Contents 3.1 Determining the Status of Reptile Species and Populations................................................... 48 3.2 Factors Contributing to Reptile Declines................................................................................ 49 3.2.1 Habitat Loss................................................................................................................. 49 3.2.2 Unsustainable Removal............................................................................................... 50 3.2.3 Environmental Contamination.................................................................................... 52 3.2.4 Climate Change........................................................................................................... 53 3.2.5 Invasive Species........................................................................................................... 53 3.2.6 Disease and Parasitism................................................................................................ 55 3.2.7 Cascading Declines..................................................................................................... 55 3.3 Global Status of Reptile Populations....................................................................................... 56 3.3.1 Testudines.................................................................................................................... 56 3.3.2 Crocodilians................................................................................................................. 57 3.3.3 Squamates: Lacertilians.............................................................................................. 58 3.3.4 Squamates: Serpents.................................................................................................... 59 3.4 Conclusion...............................................................................................................................60 Acknowledgments............................................................................................................................. 61 References......................................................................................................................................... 61 Reptiles have been considered by some to be of “minor importance,” and their disappearance has been suggested to “not make much difference one way or the other” (Zim and Smith 1953). Linnaeus himself described reptiles in his 1758 Systema Naturae as “foul and loathsome animals … abhorrent because of their cold body … fierce aspect … and squalid habitation.” Thankfully, such sentiments are increasingly outdated as scientists reveal the significant roles that reptiles play in many ecosystems. Although reptiles remain among the least studied vertebrate groups and are still frequently considered of less general interest than other fauna (Gibbons 1988; Bonnet et al. 2002), interest in the preservation of biodiversity, and consequently interest in reptile conservation, is growing. Declines of reptile populations, whether unnoticed or widely documented, are troubling not just because of the ecological relevance of reptiles to many habitats, but also because they portend a general decay of environmental health similar to declines of other species. Regardless of the motivation, the desire to conserve reptiles and to better understand their ecology requires knowledge of their status, distribution, and the factors that contribute to their decline. Our goals in this chapter are to describe the major anthropogenic threats faced by reptiles and to summarize the perceived global conservation status of the traditionally recognized major taxonomic groups of reptiles. Specifically, we discuss the Crocodilia, Squamata, and Testudines, all historically included in the class Reptilia. Unfortunately, because so little is known about the basic ecology, distribution, and status of amphisbaenids, we exclude discussion of them from this chapter. Likewise, 47

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given the low diversity (2 species) and extreme geographical restriction of tuataras (Sphenodontia), we do not separately review this group. Our specific objectives in this chapter are to

1) call attention to the lack of data on the status of reptile populations and describe strategies for documenting range-wide and localized declines, 2) describe the major threats facing reptile populations globally, and 3) describe current patterns of imperilment in reptile populations.

Two major reviews have recently described the main threats to reptile populations, and we expand on these using recent examples from an ever-growing body of literature (Gibbons et al. 2000; Irwin and Irwin 2005).

3.1 Determining the Status of Reptile Species and Populations As with many fauna in recent years, conservation biologists have raised concerns over reported reptile declines, some of which have garnered widespread recognition (e.g., Asian turtle crisis; Buhlmann et al. 2002). Nevertheless, the considerable and notable lack of information regarding the status of reptiles in most regions has hampered full understanding and appreciation of their current plight. Without question, some reptile populations have been extirpated and numbers of some species have declined with little indication of the underlying cause. In some cases, concerns about declines are based wholly on anecdotal observations or on a growing perception of a species’ rarity without accompanying quantitative data. Clearly, a primary goal of herpetologists and wildlife biologists should be to clarify the global status and distribution of reptile populations. The World Conservation Union (IUCN) has been a global leader in assessing the status of many floral and faunal species as part of its ongoing Red List program (Baillie et al. 2004). Although the IUCN has comprehensively assessed birds, mammals, and amphibians to date, a global reptile assessment was only recently begun. Currently, reptiles remain one of the least known vertebrate taxa, and the conservation status of only about 6% of species has been assessed (Baillie et al. 2004). Some impediments to determining the status of reptiles derive from difficulties in studying or monitoring their populations. For example, many reptiles are characterized by cryptic coloration or behavior, which can impede observation or capture (Zug et al. 2001; Dorcas and Willson 2009). Detectability of reptiles also depends to some degree on the survey technique used, the seasonal or daily timing of surveys, and the environmental conditions during which surveys are conducted (Todd et al. 2007). Additionally, the status of reptiles can be assessed on at least 2 scales: local populations and regional distributions. Concerns over declines of some reptiles arise from general impressions of range-wide or regional contractions in species’ distributions. Obviously, such declines are cause for alarm because they may point to wide-ranging or systemic threats to species. However, adequate quantitative documentation of range-wide or regional declines in reptile species is infrequent, possibly due to the time and collaboration required to collect sufficient data at such large spatial scales. Nevertheless, successful models of range-wide investigations do exist and demonstrate the usefulness of constructing large-scale overviews of changes in species’ distributions. In 1 example, investigators compiled distributional information from published scientific records, unpublished reptile surveys, museum databases, state natural heritage programs, and contributions by individual researchers to compare the historical vs. currently known distribution of the southern hognose snake, Heterodon simus (Tuberville et al. 2000). This collaborative effort revealed substantial declines of the species in portions of its range and provides a powerful example for other studies of distributional change in reptiles. In some cases, resurveying habitats across a species’ range can also reveal that populations may be more widespread than previously suspected, as was the case for the sharp-tailed snake, Contia tenuis, in Oregon (Hoyer et al. 2006).

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At a finer resolution, studies of individual reptile populations are also useful for revealing the plight of species, albeit at a more limited scope. As pointed out by other authors, studies of individual populations should use standardized methods to monitor changes in abundance or density over long periods of time (Gibbons et al. 2000). As an example, a long-term study of the abundance of eastern kingsnakes, Lampropeltis getula, revealed an alarming disappearance of the species from 1 site over 30 years (Winne et al. 2007). Although the decline could not be attributed to any single cause in this case, the examination of local populations is often more likely to reveal specific causes for a species’ decline than are range-wide studies. Ultimately, a combination of studies at multiple scales will provide the most comprehensive assessment of the status and distribution of reptiles, as has been the case with the Texas horned lizard, Phrynosoma cornutum (Donaldson et al. 1994). Lastly, as reptile populations fall under increasing scrutiny by scientists and conservation managers, it will be important to distinguish between natural declines and anthropogenic ones, and even to determine whether fluctuations in distribution or abundance represent “declines” per se (Alford and Richards 1999; Pechmann 2003). All animal populations presumably experience some level of normal fluctuation in abundance that will vary depending on the species or population in question. Thus, short-term monitoring that provides limited snapshots of population size may reveal current status but will not expose longer-term population trends or their causes (Gibbons et al. 2000). For this reason, the value of long-term studies and the data they generate cannot be overstated. Nevertheless, accumulation of numerous accounts from short-term studies may reveal a declining trajectory that can lend credence to conservation concerns for a given species or population (Gibbons et al. 2000). The scientific community would do well to take notice of the incredible mobilization of inventory and monitoring programs and other research activities that have followed recognition of the acute imperilment of many amphibians. Proactive recognition of the need to closely monitor reptiles may be instrumental in preventing or mediating their declines and minimizing the economic cost of reactive, and sometimes belated, conservation efforts.

3.2  Factors Contributing to Reptile Declines Establishing a causal link between any specific factor and declines of reptile populations can be difficult but is of foremost concern for effective conservation. Although in some cases 1 factor may weigh heavily on a population, multiple interacting factors nearly always affect a species’ abundance and distribution. Several factors have been identified as threats to reptile populations and are implicated in declines of at least some reptiles, including habitat loss and fragmentation, unsustainable removal, anthropogenic environmental contamination, climate change, invasive species, disease and parasitism, and trophic cascades, and we discuss these in detail in the following sections. Two other seldom mentioned but nonetheless important factors bearing on the status of reptile populations are social apathy and special or political interests. Indeed, social apathy can be a major obstacle to reptile conservation because many reptiles are subjects of personal derision, a problem that must be overcome before appropriate motivation can spur conservation interest (Gibbons 1988). Similarly, the willingness of nongovernmental organizations and state, provincial, or national governments to recognize the plight of declining species and the need for conservation effort often depends on special or political interests and will undeniably have considerable impact on the persistence of many reptiles.

3.2.1  Habitat Loss Habitat loss, including degradation, fragmentation, or conversion for other use, is typically regarded as the single greatest cause of faunal declines globally (Wilcove et al. 1998; Sala et al. 2000). Thus, the fact that habitat loss is considered to be the leading cause of reptile declines is not surprising (Mittermeier et al. 1992; Gardner et al. 2007). Habitat loss due to conversion of land for human use typically occurs for agriculture, housing or infrastructure development, commercial forestry, and to support recreation, including constructing golf courses or dredging lakes and other aquatic habitats.

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Habitat loss from conversion can result in direct animal mortality that is often difficult to quantify. Nevertheless, mortality of turtles and aquatic snakes during lake dredging has been documented (Aresco and Gunzburger 2004), and entombment of live gopher tortoises (Gopherus polyphemus) during land development was widespread in Florida prior to a ban on the practice (Cox 2007). Ultimately, however, immediate mortality from habitat alteration likely poses less threat than the subsequent longterm, indirect effects of habitat loss and degradation on survival and reproduction. Habitat loss can affect reptiles indirectly by limiting their ability to meet ecological needs for survival and reproduction. For example, many reptiles decline in abundance over time following the clearing of primary forest or conversion to plantation forest (Glor et al. 2001; Kanowski et al. 2006). In fact, the decline of multiple reptile species in the southeastern United States has followed widespread and nearly complete loss of native longleaf pine habitat (Ware et al. 1993; Gibbons et al. 2000). At a finer scale, 1 study from the southeastern United States demonstrated that planted pine forests and recent clear-cuts supported reduced abundances of small snakes compared with open-canopied partially harvested forests (Todd and Andrews 2008). The precise mechanisms of decline remain unknown but are presumably related to a general degradation of habitat quality from anthropogenic land conversion. Additional examples of the effects of habitat degradation include the loss of foraging and refuge due to bush-rock collection, which has contributed to the decline of the Australian broad-headed snake (Hoplocephalus bungaroides; Shine et al. 1998), and reductions in lizard abundance due to human-induced bush encroachment in Africa (Meik et al. 2002). Loss and degradation of aquatic habitats also pose a serious threat to reptiles. Notable examples include declines of the crocodilian fauna of the Ganges and Yangtze Rivers (IUCN 2009), which have become increasingly imperiled following damming, flow modification, and general degradation of river habitat (Dudgeon et al. 2006). Similarly, channelization and dam building have been implicated in declines of river-dwelling North American map turtles (Graptemys spp.; Kofron 1991). The vulnerability of sea turtles to coastal development that degrades or eliminates nesting habitat has been appreciated for decades (Lutcavage et al. 1997; Spotila 2004). Likewise, the effects of terrestrial habitat alteration that disturbs or eliminates nesting and refuge sites of freshwater turtles can also be severe (Buhlmann 1995; Burke and Gibbons 1995). Many semiaquatic snake species that use wetland habitats share the plight of turtles if wetlands are lost or terrestrial habitat around sensitive aquatic resources is altered (Roe et al. 2003, 2004; Willson et al. 2006). Habitat fragmentation is the emergence of discontinuities in an organism’s preferred environment. Habitat fragmentation may occur due to natural processes but increasingly results from anthropogenic habitat loss or land conversion that isolates remaining patches of suitable habitat. The degree to which habitat fragmentation threatens a species depends on how greatly a species’ movements are affected by the interspersed barriers that separate remaining usable habitat. Some lizard, snake, and turtle species are vulnerable to habitat fragmentation, with general declines in abundance being reported (Dodd 1990; Kjoss and Litvaitis 2001; Driscoll 2004), whereas others are not (Driscoll 2004). In other cases, habitat fragmentation affects the demography of remaining reptile populations. For example, patch size was positively correlated with abundance, survivorship, and recruitment of Florida scrub lizards, Sceloporus woodi (Hokit and Branch 2003). Of particular concern is the possible role that roadways play in fragmenting habitat. Because road mortality of turtles and snakes is often high (Aresco 2003; Gibbs and Steen 2005; Andrews and Gibbons 2008; see the next section), roads effectively become barriers that separate and isolate habitat (Roe et al. 2006). Again, however, species will differ in the extent to which roads act as barriers to movement (Andrews and Gibbons 2005), and therefore fragment populations.

3.2.2  Unsustainable Removal Removal of reptiles from wild populations occurs both commercially and noncommercially for food, “traditional” medicine, curios, and the pet trade, as well as unintentionally as by-catch in

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other harvesting activities and, increasingly, as a result of road mortality. Although removal per se is not necessarily harmful to population persistence — many reptile populations could presumably sustain some low level of harvest — removal at unsustainable rates is a serious threat that places many reptile populations and species in peril. To date, few studies have demonstrated sustainability of removal activities on reptile populations. In contrast, studies documenting intense levels of reptile harvesting and subsequent declines in wild populations are common. Perhaps the mostly widely recognized removal-driven peril results from the ongoing use of reptiles for food, skins, or “traditional” medicines. Imperilment of Asian turtles due to unsustainable removal has reached crisis levels and has grim consequences for the persistence of many freshwater and terrestrial turtles if not remedied (Buhlmann et al. 2002). The exploitation of turtles, however, is not restricted to Asia; many Central and South American cultures relish turtles and their eggs, resulting in continued threats to both freshwater and marine turtle populations (Lagueux 1991; Thorbjarnarson et al. 1997). Moreover, the consumptive use of reptiles is not limited to turtles. In Asia, snakes face rapidly growing pressure from exploitive use, with as many as 1 million snakes being harvested in northeast China and nearly 8 million kg traded each year across the country (Zhou and Jiang 2004, 2005). Likewise, in Cambodia, an estimated 6.9 million aquatic snakes are removed annually from Tonle Sap Lake to feed the growing crocodile farms in that region (Brooks et al. 2007). Subsequently, hunters have reported a 74% to 84% decline in snake catch from 2000 to 2005 (Brooks et al. 2007). Pythons (Python spp.), too, face significant harvest pressure in many parts of Indonesia (Shine et al. 1999). Monitors (Varanus spp.) and tegus (Tupinambis spp.) are heavily harvested for their skins at rates of as much as 1 million animals per year in the case of South American tegus (Pianka and Vitt 2003; Mieres and Fitzgerald 2006). In Africa, bushmeat consumption often extends to highly endangered species such as the dwarf crocodile, Osteolaemus tetraspis (Willcox and Nambu 2007). The complete list of affected species is lengthy (see also reviews in Gibbons et al. 2000; Irwin and Irwin 2005) and a precautionary policy of preventing massive exploitation until sustainable removal limits are identified appears to be the best method of ensuring the persistence of reptile populations. The commercial removal of reptiles from wild populations for use as pets is another consumptive use affecting reptiles globally. In many cases, evidence linking collection of animals for the pet trade to declines of wild reptile populations is lacking because long-term studies of the status of reptile populations following targeted collecting have not been conducted. Nevertheless, collection of reptiles for the pet trade may endanger wild populations due to the large scale at which some reptiles are collected. Many species of turtles, snakes, chameleons, and other lizards are under increasing demand, and collection in wild populations continues at high levels (Reed and Gibbons 2003; Carpenter et al. 2004; Schlaepfer et al. 2005) that can threaten population persistence (Webb et al. 2002). Removal of individual reptiles also occurs as by-catch in fisheries and from intentional and unintentional road mortality. Mortality of long-lived sea turtles in longline fisheries and shrimp nets is likely unsustainable and contributes to the rapid and ongoing declines of many sea turtle species (Lewison et al. 2003, 2004). Similarly, both commercial and recreational crab trapping have been implicated in declines of the North American diamondback terrapin, Malaclemys terrapin (Bishop 1983; Dorcas et al. 2007), a species already heavily depleted from commercial exploitation for food in the 1800s to early 1900s (Carr 1952). In developed countries, roads represent an additional, substantial, and often ignored source of mortality in many reptile populations. Snakes and turtles are probably hardest hit by road mortality because they are large, often move slowly, and are sometimes direct targets for persecution by motorists (Andrews and Gibbons 2005; Ashley et al. 2007). Increasingly male-biased sex ratios in turtle populations attest to the potential long-term effects of road mortality on reptiles; female turtles are more likely to cross roads during nesting forays and are enticed to oviposit on road shoulders due to the presence of sunny nesting habitat, leading to disproportionately higher female mortality (Aresco 2005; Steen et al. 2006).

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3.2.3  Environmental Contamination Release of contaminants — including pesticides, herbicides, heavy metals, and radioactive waste — into the environment has been listed as 1 of the 6 major contributors to the global decline of reptiles (Gibbons et al. 2000). Reptiles exhibit a suite of ecological and life history characteristics that make them particularly vulnerable to contaminants (Hopkins 2000). With the exception of a few lizard and turtle species, reptiles are strictly carnivorous, and many occupy high trophic positions within food webs. Thus, reptiles are at risk from biomagnification of contaminants. Additionally, many reptiles are long-lived and have small home ranges compared to similar-sized endotherms, making them susceptible to long-term contaminant exposure and subsequent bioaccumulation (Hopkins 2000; Shelby and Mendonca 2001; Bergeron et al. 2007). Although reptiles may be at particularly high risk from contaminants, they are currently the least studied vertebrate group in ecotoxicology (Hopkins 2000; Chapter 1, this volume). Within reptiles, ecotoxicological research has primarily been restricted to turtles and crocodilians, and our knowledge of the effects of toxins on squamates is still limited (Campbell and Campbell 2001). The most overt measurable effect that contaminants can have on reptiles is direct mortality of individuals resulting from exposure. Several studies have reported mortality of reptiles in association with intentional (e.g., pesticide application) or accidental (e.g., spills or contaminant leakage) introduction of toxins into the environment (reviewed in Campbell and Campbell 2000, 2001), but few authors have related such acute mortality events to population declines. However, Romero and Wilkelski (2002) noted population declines of Galapagos marine iguanas (Amblyrhynchus cristatus) following a low-level oil spill. Declines were not observed on islands that remained unaffected by the spill. Romero and Wilkelski (2002) speculated that iguanas died of starvation after the digestive bacteria in their guts were killed by oil residues found in their diet of marine algae. In a similar example, Ernst et al. (1994) noted the disappearance of yellow-blotched map turtles (Graptemys flavimaculata) from sections of river immediately downstream from a paper mill. However, it is unclear if changes in turtle abundance represented population declines or movement of turtles out of contaminated areas due to lack of prey (aquatic invertebrates) or other factors. To date, reports of population declines associated with environmental contamination are largely restricted to turtles and large lizards. Cryptic behavior and/or low activity levels (which result in low capture rates during surveys; Dorcas and Willson in press) of many snakes and smaller lizards would conceal mortality events and hamper detection of population declines. Although less obvious than direct mortality, sublethal effects of contaminants may be more detrimental to the long-term persistence of reptile populations. High tissue loads of various contaminants have been documented from reptiles in the field (e.g., lizards and snakes; reviewed in Campbell and Campbell 2000, 2001; Bergeron et al. 2007; this volume), and sublethal effects of contaminants on reptile locomotor performance (e.g., Hopkins et al. 2005; Holem et al. 2006; Hopkins and Winne 2006; DuRant et al. 2007a) and metabolic energy consumption (e.g., Hopkins et al. 2005; DuRant et al. 2007b) have been demonstrated in the laboratory. Although these studies provide insight on the mechanisms linking sublethal contaminant exposure to population dynamics, few studies have attributed declining reptile populations to sublethal contaminate exposure. A well-cited exception is the decline of American alligators (Alligator mississippiensis) in a Florida lake contaminated with estrogenic compounds (Guillette et al. 1994; Semenza et al. 1997). Alligator declines were attributed to reproductive failure resulting from reduced testosterone levels and gonadal malformations. Likewise, Shelby and Mendonca (2001) found reduced testosterone levels in some male yellowblotched map turtles (Graptemys flavimaculata) from polluted habitats, suggesting that effects of pollutants on reproduction may have been responsible for population declines previously observed at that site. Although much of the investigation of indirect effects of contaminants on reptiles has focused on the effects of endocrine disrupters on reproduction, exposure to contaminants may also affect energy acquisition and expenditure. For example, Hopkins et al. (1999) found that banded watersnakes (Nerodia fasciata) collected from a wetland polluted with coal combustion waste had

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elevated tissue concentrations of trace metals and standard metabolic rates that were 32% higher than those of snakes from an unpolluted site. Elevated metabolic rates presumably limit energy availability for growth and reproduction in snakes from contaminated sites. Finally, consider possible synergistic effects of contaminants with other threats facing reptiles. For example, a plausible scenario is one in which sublethal exposure to contaminants compromises immunocompetence, resulting in outbreaks of opportunistic pathogens that might otherwise be benign or manageable under normal circumstances. Although these types of questions have not been addressed in reptiles, consideration of the effects of multiple stressors on reptiles will undoubtedly become increasingly important as human populations continue to grow and expand around the globe.

3.2.4  Climate Change Global climate change has been an ongoing process throughout the evolutionary history of reptiles and has been nearly continuous in the past 65 million years (Zachos et al. 2001). However, given indications that recent global climate warming is occurring at a pace unprecedented in recent history (IPCC 2007), rapid climate change is particularly relevant in our consideration of threats to reptile populations. Possible effects from climate change fall broadly into categories of direct and indirect effects, both of which have either caused or are expected to cause changes in reptile populations. Because reptiles are ectothermic, they are highly dependent on suitable external temperatures to regulate their own body temperatures and support metabolic and other functions. Subsequently, direct effects of climate change may manifest as changes in growth rates or the age at onset of reproductive maturity, as shown in painted turtles, Chrysemys picta (Frazer et al. 1993). Theoretical models further suggest that changes in global climate can have profound effects on reptile populations. Dunham (1993) used individual-based models to estimate physiological responses of Big Bend Canyon Lizards (Sceloporus merriami) to climate change and predicted that increases in air temperature of 2 °C to 5 °C could constrict activity sufficiently to drive populations to extinction. Others have expressed concern that climate change may have a considerable impact on reptiles with temperature-dependent sex determination, such as some turtles, lizards, and crocodilians, by altering sex ratios within populations (Janzen 1994). Other obvious effects that global climate change may have on reptile populations include direct and indirect influences on habitat suitability. For example, changes in temperature and precipitation may directly affect the habitability of a reptile’s environment and could cause shifts in reptile distributions at a large scale. Araújo et al. (2006) explored possible scenarios of habitat change in Europe and concluded that, although suitable habitat for European reptiles is likely to expand under most circumstances, limited dispersal abilities of reptiles may increase their vulnerability to climate change. Dispersal capability may be further constrained by the ever-increasing habitat fragmentation that accompanies human population growth. Furthermore, regional changes in precipitation and temperature regimes are likely to broadly affect community composition, and some landscapes may change dramatically (Guertin et al 1997; Still et al. 1999). Whitfield et al. (2007) suggested 1 instance of indirect effects of climate change when they documented a steady decline in several lizard species at La Selva, Costa Rica, over a 35-year period. They attributed declines to climatedriven reductions in the quantity of leaf litter. Also, severe climatic events are expected to become more frequent due to the destabilization of regional weather patterns under many global warming scenarios (IPCC 2007). Subsequently, droughts or other meteorological events such as cyclones may negatively affect reptile populations (Seigel et al. 1995; Willson et al. 2006).

3.2.5  Invasive Species Recent expansion of human populations and increases in global transportation and trade have resulted in introduction and establishment of many species in areas outside of their native geographic range. Many introduced species have subsequently proliferated, resulting in severe ecological and

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economic damage (Pimentel et al. 2000). Consequently, invasive exotic species are currently recognized as one of the foremost threats to global biodiversity (Park 2004), including reptiles (Gibbons et al. 2000). Invasive exotic species affect native species in a variety of ways. One of the most obvious ways introduced animals can affect reptile populations is by directly preying upon them. Predator introductions that have resulted in the decline or extirpation of many species are especially obvious on island ecosystems that were previously bereft of predators. For example, introduced predators (mongoose, Herpestes javanicus, and rats, Rattus rattus) have been identified as the single greatest threat to snakes on the Lesser Antilles and have been implicated in at least 6 historical snake extirpations and at least 1 historical extinction in that region (Henderson 2004). Similarly, introduction of brown treesnakes (Boiga irregularis) to the island nation of Guam has devastated the vertebrate fauna of that island, including populations of several native lizards (Fritz and Rodda 1998). A final example demonstrates that introduced predators need not be large vertebrates. Since their accidental introduction to Mobile, Alabama, in the 1930s, imported red fire ants (Solenopsis invicta) have spread throughout much of the southeastern United States (Wojcik et al. 2001). Mount (1981) expressed concern about the potential impacts of predation by S. invicta on vertebrates in the Southeast, noting observations of fire ant predation on eggs and hatchlings of several reptile species. He also made anecdotal observations of declines of many litter-dwelling snakes and lizards, and large terrestrial oviparous snakes in the Alabama coastal plain in conjunction with S. invicta invasion. Although quantitative evidence of the effects of S. invicta on native reptiles has been slow to emerge, fire ants are documented predators of turtle nests (Buhlmann and Coffman 2001) and have been implicated in the decline of southern hognose snakes (Heterodon simus; Tuberville et al. 2000), eastern kingsnakes (Lampropeltis getula; Wojcik et al. 2001, Winne et al. 2007), and Texas horned lizards (Phrynosoma cornutum; Goin 1992). Importantly, the effects of S. invicta on reptiles may be exacerbated by habitat disturbance, possibly leading to synergistic effects of habitat alteration and predation (Todd et al. 2008). Introduction of exotic prey can also have profound effects on reptile populations. In some cases, exotic prey may possess defenses (e.g., poisons, morphological defenses) to which native predators are unaccustomed, resulting in direct mortality of reptiles that attempt to consume the exotics. For example, exotic cane toads (Bufo marinus) possess potent parotoid secretions and have become abundant in many areas of tropical Australia since their introduction in 1929 (Lampo and de Leo 1998). Covacevich and Archer (1975) noted several instances of direct mortality of snakes and monitor lizards (Varanus spp.) that attempted to ingest the toads. Other authors have observed declines of several snake and lizard species following arrival of B. marinus (Phillips et al. 2003), as well as mortality of Australian freshwater crocodiles. Moreover, laboratory studies have shown that many Australian snake species are sufficiently vulnerable to toad toxins to die after ingesting a single toad, prompting Phillips et al. (2003) to suggest that the toads threaten as much as 30% of Australia’s terrestrial snake species. In other cases, exotic prey may be palatable, but of poorer nutritional quality than native prey taxa. For example, exotic Argentine ants (Linepithema humile) have been introduced worldwide (Suarez et al. 2001) and eliminate nearly all native ground-dwelling ants when they invade new habitats (Suarez et al. 1998). Suarez and Case (2002) demonstrated that Argentine ants represent an inferior prey resource for coastal horned lizards (Phrynosoma coronatum), a species that has declined dramatically in California. They found that hatchling P. coronatum fed a diet of introduced L. humile exhibited zero or negative growth, but resumed normal growth when switched back to a diet of native ants. Introduced exotic prey taxa may sometimes be beneficial to native reptiles. For example, round gobies (Neogobius melanostomus) introduced into the Great Lakes region of North America have become favored prey of the federally listed Lake Erie watersnake (Nerodia sipedon insularum), resulting in increased growth and body sizes of snakes (King et al. 2006a). Introduction of gobies has been implicated as a partial cause for the recovery of this snake in recent years (King et al. 2006a) to levels that warrant delisting under the US Endangered Species Act (King et al. 2006b). Introduced species may also exert substantial indirect effects on native reptile populations through competition for resources. For example, introduced geckos (Hemidactylus frenatus) compete with and

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have displaced native geckos (Lepidodactylus lugubris) throughout the tropical Pacific (Case et al. 1994). Likewise, competition with H. frenatus has caused declines and extirpations of several species of night geckos (Nactus spp.) on the Mascarene Islands. Introduced slider turtles (Trachemys scripta) compete with native European pond turtles (Emys orbicularis) for basking sites (Cadi and Joly 2003), resulting in weight loss and reduced survival of E. orbicularis in experimental mixed populations compared with controls (Cadi and Joly 2004). Importantly, in some cases, introduced species may gain a competitive advantage because of release from their native pathogens or parasites (Reed 2005). Another indirect means by which invasive species can affect native reptiles is through habitat modification. The most obvious examples of this phenomenon occur in cases where invasive plant species displace native vegetation, rendering habitat unsuitable for native species. For example, lush growth of invasive annual plants in the Mojave Desert of the American Southwest negatively affects desert tortoises (Gopherus agassizii), primarily through increased fire frequency (Brooks and Pike 2001). Likewise, habitats dominated by exotic rubber vine (Cryptostegia grandiflora) are avoided by native Australian lizards (Valentine 2006). Finally, introduced species may be important vectors for disease and parasites (see Section 3.2.6). Unlike other ways in which exotic species may affect reptiles, exotic species do not need to become established in the wild to serve as disease vectors. In fact, release of a single infected individual, or human contact with a wild reptile after handling an infected captive reptile, may be sufficient to introduce a pathogen to native reptile populations. Reed (2005) cautioned that release of captive boas and pythons could be an important source of disease to native snakes such as the federally threatened eastern indigo snake (Drymarchon couperi) or boid species native to the United States (rubber boa, Charina bottae, and rosy boa, Lichanura trivirgata).

3.2.6  Disease and Parasitism Pathogens and parasites have long been recognized as potentially important factors regulating natural populations (Anderson and May 1978; Dobson and Hudson 1986). Virtually every species hosts a multitude of parasites and pathogens, some of which can cause dramatic population fluctuations (e.g., Hudson et al. 1998). However, when human activities alter rates of disease transmission or reduce resistance of animals to disease, the results can be catastrophic (Daszak et al. 2000). For example, outbreaks of pathogenic chytrid fungus (Batrachochytrium dendrobatidis) have devastated amphibian populations worldwide (Daszak et al. 2003). Moreover, the spread of chytrid appears to have been facilitated by global climate change (Pounds et al. 2006), introduction of exotic species (Mazzoni et al. 2003), and direct spread by humans (Weldon et al. 2004). Disease outbreaks have been implicated in declines of several reptile taxa and are apparently of particular concern for turtles. For example, incidence of upper respiratory tract disease (URTD) infections of gopher tortoises (G. polyphemus) and desert tortoises (G. agassizii) has increased in recent years, and URTD has been implicated in declines of some populations (Dodd and Seigel 1991; Seigel et al. 2003). Upper respiratory tract disease may have been introduced into natural populations through release of infected captive individuals (Dodd and Seigel 1991), and authors have expressed concern that spread of this disease could be exacerbated through translocation of infected animals for conservation purposes (Dodd and Seigel 1991). Disease outbreaks have been noted in other protected turtles, such as the flattened musk turtle (Sternotherus depressus; Dodd 1988; Fonnesbeck and Dodd 2003) and green sea turtle (Chelonia mydas; Chaloupka and Balazs 2005); however, factors underlying these outbreaks are poorly understood.

3.2.7  Cascading Declines An additional consideration seldom explicitly addressed is how reptiles respond to declines in other taxa (but see Irwin and Irwin 2005). Although ecologists have yet to form a consensus about the role of biodiversity in maintaining ecosystem function (Thompson and Starzomski 2007), loss of

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important species could result in subsequent extinction of previously unaffected taxa. Living in a time when species extinctions are increasing at an alarming rate, we are probably only beginning to see the long-term effects of extinctions that have already occurred. An obvious concern regarding cascading declines is the possible response of reptiles to the recent catastrophic declines of many amphibian populations. Many snake species are specialists that feed exclusively on amphibians (Toledo et al. 2007); we might therefore expect that these species would experience declines concomitant with those of amphibians. Whiles et al. (2006) reported the disappearance of several common frog-eating riparian snakes soon after amphibian declines related to chytridiomycosis in Panama. Similarly, in the Sierra Nevada Mountains of the American West, the presence of mountain garter snakes (Thamnophis elegans) is strongly associated with the presence of anurans (Matthews et al. 2002). Matthews et al. (2002) suggest that amphibian declines associated with stocking of predatory trout may have a strong adverse effect on populations of T. elegans. Cascading declines may be most frequently associated with the disappearance of keystone species, those species that play a role disproportionate to their abundance in maintaining community composition and ecosystem function (Power et al. 1996). For example, the gopher tortoise (G. polyphemus) is considered a keystone species (Eisenberg 1983) because its burrows provide critical refugia for a variety of upland species in the southeastern United States. Gopher tortoises have declined across their range (McCoy et al. 2006), and loss of tortoise burrows is considered a serious threat to persistence of the federally threatened eastern indigo snake (Drymarchon couperi; Stevenson et al. 2003) and eastern diamondback rattlesnake (Crotalus adamanteus; Timmerman and Martin 2003), among other species. Although such cascading declines affect all species, reptiles, which often have highly specialized food and habitat requirements, may be less able than more generalist taxa to withstand sequential removal of individual species from ecosystems.

3.3 Global Status of Reptile Populations As discussed previously, the status of many reptiles remains unknown and is the subject of ongoing global assessment by the IUCN (Figure 3.1). Here, we describe the current perceived status of reptile populations and provide overviews of reptile declines among turtles, crocodilians, lizards, and snakes. Our reports on the current status of reptile populations represent our best understanding of the available scientific literature and current information from the 2008 IUCN Red List (IUCN 2009).

3.3.1  Testudines A greater proportion of turtles are recognized as imperiled and in categories of conservation concern than any other group of reptiles excluding the 2 tuatara species, Sphenodontia (Figure 3.1). Overall, 42% of turtle species included in the IUCN 2004 global assessment were classified as threatened (including all IUCN categories of imperilment; Baillie et al. 2004). However, at that time only a portion of all described species were evaluated. Consequently, the actual rate of imperilment of evaluated species is closer to 62% (Baillie et al. 2004). Perhaps most alarming is that the IUCN currently lists a total of 8 turtle species that have gone extinct in the wild in modern times (IUCN 2009). The Turtle Conservation Fund even compiled a list of 25 additional turtle species on “death row,” that is, the most endangered species of tortoises and freshwater turtles in the world (Buhlmann et al. 2002). Although not included in the list, all 7 species of sea turtles (families Cheloniidae and Dermochelyidae) are considered imperiled. The taxonomic distribution of endangered species (critically endangered [CE] and endangered [EN]) includes species from each of the 11 families of freshwater turtles, the single family of tortoises, and both families of sea turtles (Baillie et al. 2004). Turtles are long-lived (Gibbons 1987), and commercial harvesting of wild populations of most species is not sustainable (Reed and Gibbons 2003). Nonetheless, turtles differ from other reptile groups in that human consumption is the documented cause for the majority of declines on a global scale (Buhlmann et al. 2002). Commercial harvesting in Southeast Asia is a major cause

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The Global Status of Reptiles and Causes of Their Decline 23

2

4450

2900

280

Percentage of Species

100%

Unknown

75%

Least Concern

50%

Threatened

25% 0%

Extinct Crocodilia

Sphenodontia SquamataLacertilia

SquamataSerpentes

Testudines

Figure 3.1  Status of the major lineages of reptiles according to the World Conservation Union (IUCN) Red List in 2009. Status categories include “extinct” (including species extinct in the wild but extant in captivity), “threatened” (including IUCN categories “critically endangered,” “endangered,” and “vulnerable”), “least concern” (including IUCN categories “lower risk” and “near threatened”), and “unknown” (including species that have not been evaluated by the IUCN and those that have been evaluated but were deemed “data deficient”). Numbers above bars indicate the approximate number of species within each lineage according to Zug et al. (2001). Data were accessed from the IUCN database on March 18, 2009 (IUCN 2009). Note that although bars are difficult to see, 11 species of lizards (0.25%) and 3 species of snakes (0.10%) are listed as extinct by the IUCN as of 2009.

for concern, with some turtle species clearly on a trajectory toward extinction at current rates of removal from the wild (Baillie et al. 2004). Indeed, of 73 species of tortoises and freshwater turtles classified as endangered and critically endangered in 2002, more than half were from Asia, with the remaining species being distributed geographically among North America, Mesoamerica, South America, the Mediterranean, Africa, and Australasia (Buhlmann et al. 2002). Outside of Asia, terrestrial terrapins and tortoises are imperiled by excessive harvesting combined with loss of suitable habitat (Baillie et al. 2004). Collection and removal of turtles from North America, mostly for the export trade, has also been significant (Franke and Telecky 2001; Ceballos and Fitzgerald 2004) and could understandably be implicated in the decline of some species in the wild. Sea turtles, all of which are classified as imperiled, are subject to unique threats, with declines attributed to mortality from incidental by-catch, harvesting of turtles and eggs for consumption, and degradation of nesting and foraging habitat (Lutcavage et al. 1997; Spotila 2004).

3.3.2  Crocodilians A quarter of a century ago, every crocodilian species in the world was categorized as endangered or threatened. Ironically, because the fate of only 7 of the 23 species remains uncertain at the beginning of the 21st century, the group is considered by many conservationists to be a major success story. Conservation efforts by the Crocodile Specialist Group of the Species Survival Commission–IUCN are generally viewed as the cause for an upturn in the status of two-thirds of the crocodilian species that have traditionally suffered from the pressures of harvesting and habitat loss (IUCN 2009). Despite the noted conservation successes in reducing declines and extinction threats for some crocodilian species, the status of 7 species, including the Chinese alligator (Alligator sinensis), the black caiman (Melanosuchus niger), the Indian gharial (Gavialis gangeticus), and 4 species of crocodile, remains one of teetering on the brink of extinction. In mainland Asia, only about 150 Chinese alligators are estimated to be present in their native habitat, and the Siamese crocodile (Crocodylus siamensis) is effectively extirpated from its native home of Thailand (formerly Siam), with few populations persisting in other parts of Southeast Asia. Only a single established population of the

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Philippine crocodile (Crocodylus mindorensis) is known to exist at this time, with the species now occupying less than 20% of its former range of most of the Philippines Archipelago. Likewise, the Indian gharial is considered to be critically endangered despite strong and effective conservation programs in some parts of the country. In fact, nearly 80 gharials, representing 6% of the known population, died from unknown causes in an Indian forest preserve from December 2007 to January 2008 (Mahmood 2008). Initial reports identifying liver cirrhosis in the dead animals suggested that parasites or environmental contaminants may have played a role in the deaths. In the western hemisphere, no fewer than 3 crocodilian species remain on the list of special concern, although some seem to be recovering through regional conservation efforts. In tropical South America, the most endangered species are the black caiman and the Orinoco crocodile (Crocodylus intermedius), a species whose decline has been attributed in part to the unrestricted use of pesticides for agricultural purposes. The Cuban crocodile (Crocodylus rhombifer), whose geographic range once included several islands in the West Indies, is now represented in the wild only from highly localized areas in Cuba. Nevertheless, many conservationists point to the American alligator (Alligator mississippiensis) as one of the greatest models of successful recovery for any threatened vertebrate. Most of the remaining 16 species of crocodilians — including 3 native to Africa, 2 found in Australia, and others at scattered locations in tropical America and the Pacific — are considered to be vulnerable only if the strict conservation programs in place are discontinued. Also, although many of these remaining species are not considered to be threatened with extinction throughout their geographic ranges, some still have critically endangered regional populations.

3.3.3  Squamates: Lacertilians Like many other reptiles, the status of most lizard species and populations is largely unknown (Figure 3.1). However, based on current information, lizards appear to have a small proportion of imperiled species (IUCN 2009). This is due in large part to life history attributes that make many lizards less susceptible to decline from anthropogenic factors. Notably, many lizards occur at high population densities, have short generation times, high fecundity, and are not as long-lived as other reptiles. Consequently, lizards may sometimes adjust rapidly to environmental change or rebound quickly from short-term population reductions. In fact, some lizard species fare well in humanmodified or early successional habitats (e.g., Anolis spp., Hemidactylus spp.). In several cases, these life history attributes have contributed to the successful establishment of exotic lizards introduced into areas outside of their native ranges. For example, there are more than 30 species of nonnative lizards in Florida, representing over two-thirds of the total lizard fauna in that state (Meshaka et al. 2005). Ultimately, however, several lizard species are declining and have been classified as threatened or worse under the IUCN Red List system (IUCN 2009). Moreover, the lack of data on the status of many lizard populations may further jeopardize their long-term persistence. Causes for the endangerment of lizards vary widely, but life history characteristics greatly influence the degree to which different factors threaten species. Imperiled species are those that typically have attributes such as endemism, restricted geographic ranges, large body size, long lives, late maturity, or low fecundity, which make them susceptible to population declines from anthropogenic factors. For instance, the slow-maturing, long-lived, and often endemic giant land iguanas of the Caribbean (Cyclura and Brachylophus spp.) are among the most threatened lizards globally (Pianka and Vitt 2003). In some cases, populations of Caribbean iguanas have dwindled to as few as 100 individuals. Much of their decline has been attributed to the historical harvest of these lizards for food and the introduction of nonnative pests such as mongooses, rats, goats, and pigs onto their island homes. The fate of these lizards now depends almost entirely on human intervention to preserve habitat, control introduced predators, and increase recruitment to avoid permanent extinction. Many varanoid lizards are also slow maturing and long-lived, and several of them are currently protected in parts of their ranges due to population declines (Pianka and Vitt 2003; Pianka et al. 2004). Although phylogenetically distinct from lizards, tuatara (Sphenodon spp.) share these “slow”

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life history characteristics and have long incubation periods that place them in similar jeopardy (McIntyre 1997). At least a few smaller, “typical” lizards are also threatened or critically endangered despite having high fecundity and early sexual maturity. The largest contributor to the endangerment of these species is often their endemism, restricted ranges, or highly specialized habitat requirements. High fecundity and early sexual maturity may safeguard them from declines associated with harvesting but greatly increase their risk of decline from habitat loss. In fact, habitat loss is listed as a contributing factor in the imperilment of more than half of squamates currently recognized as near threatened or worse on the IUCN Red List (IUCN 2009). For instance, the Coachella Valley fringe-toed lizard (Uma inornata), the island night lizard (Xantusia riversiana), and several island geckos from Madagascar and the Caribbean (Phelsuma spp. and Sphaerodactylus spp.) are threatened or critically endangered due to combinations of restricted geographic ranges and habitat loss. Nevertheless, some species, such as Madagascar’s Antsingy leaf chameleon (Brookesia perarmata), also presumably suffer from heavy collection, which has led to restrictions on their international trade (CITES 2003).

3.3.4  Squamates: Serpents Despite recent reports of reptile declines, the global status of snake populations has received relatively little attention. Indeed, along with lizards, snakes have yet to receive comprehensive review by the IUCN. According to the 2004 IUCN Global Species Assessment, only 3.4% of squamate species had been evaluated, compared with 67% of turtles, 90% of mammals, 100% of birds, and 100% of amphibians (Baillie et al. 2004). Five years later, the online IUCN Red List still demonstrates that the status of most squamates is unknown (Figure 3.1; IUCN 2009). Snakes are notorious for their cryptic behavior and low or sporadic activity, which seriously complicates efforts to assess population status (Parker and Plummer 1987). Thus, even for relatively well-studied species, population size or density often remains unknown (Dorcas and Willson 2009). As in lizards, risk of imperilment in snakes generally correlates more closely with life history attributes and geography than with taxonomy. Threatened species are most often those with specialized habitat requirements, small geographic ranges, or life history characteristics such as large body size, delayed sexual maturity, and/or low reproductive rates. Additionally, there is regional variation in the relative importance of threats faced by snakes such that major threats and taxa at risk vary among regions or continents. Many snakes have specialized habitat requirements, making them particularly susceptible to habitat loss or degradation. For example, many of the most threatened snake species in the eastern United States, including the eastern indigo snake (Drymarchon couperi), eastern diamondback rattlesnake (Crotalus adamanteus), pine snake (Pituophis melanoleucus), and southern hognose snake (Heterodon simus), are those associated with the nearly eliminated longleaf pine ecosystem (Todd and Andrews 2008). Likewise, loss or degradation of wetland habitats has prompted federal listing of several North American snakes, including the wetland-associated eastern massasauga (Sistrurus catenatus catenatus), San Francisco garter snake (Thamnophis sirtalis tetrataenia), and copperbelly watersnake (Nerodia erythrogaster neglecta). Small geographic range, combined with specialized habitat requirements, puts species at risk from a variety of threats to their habitat. For example, the broad-headed snake (Hoplocephalus bungaroides), considered Australia’s most endangered snake species, is restricted to small regions of rock outcrop habitat in eastern Australia and has suffered extensively from habitat degradation due to rock removal, collection for the pet trade, and canopy closure from fire suppression (Shine et al. 1998; Webb et al. 2002, 2005). Many of the most dramatic cases of imperilment due to small geographic range occur among snake species endemic to islands such as the Caribbean Lesser Antilles. The Lesser Antilles harbor 25 snake species, 87.5% of which are endemic, and are home to some of the rarest snakes in the world (e.g., the Antiguan racer, Alsophis antiguae; Daltry et al. 2001). The region has suffered between

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6 and 11 historical extirpations and at least one historical extinction, primarily due to predation by introduced mongooses (Henderson 2004). Although data are lacking, similar declines may also be occurring in other island archipelagos across the globe. Across all animal taxa, life history attributes such as large body size, delayed sexual maturity, and low reproductive output contribute to imperilment, and many of the most threatened snake species also share these characteristics (Meffe and Carroll 1997). Among snakes, body size correlates strongly with home range size (Reed and Shine 2002; Reed 2003). Thus, larger species typically need larger tracts of suitable habitat and move more extensively than smaller species, presumably putting them at greater risk from road mortality or other threats (Andrews and Gibbons 2008). Many taxonomic groups of snakes have intrinsically slow growth and low reproductive rates, making them particularly susceptible to overharvesting and less able to recover from short-term population declines. For example, many species of European and Asian vipers (Vipera spp.) are considered threatened (IUCN 2009), in part because their “slow” life history characteristics put them at risk from persecution, collection, and habitat loss. Among Australian elapid snakes, Reed and Shine (2002) found that the characteristics that correlated most strongly with species endangerment were foraging strategy (ambush foragers were most imperiled) and mating system (species with female-biased sexual size dimorphism and lacking male-male combat were frequently threatened). They postulated that snakes employing ambush foraging had more specific habitat requirements and exhibited “risky” life history attributes such as low reproductive rates and slow growth rates. Likewise, large female body size potentially increased vulnerability of females to anthropogenic sources of mortality. Their results suggest that factors contributing to endangerment in snakes may differ substantially from other taxa (especially endotherms) and are not always intuitive (Reed and Shine 2002). Unfortunately, similar macroecological analyses have not been performed for other snake groups or geographic regions. Substantial regional differences exist in the threats affecting snake populations and, consequently, in the status of snake populations. Throughout most temperate regions of Europe and North America, the paramount threat to snake populations is apparently habitat loss and degradation. However, road mortality, persecution, and collection for the pet trade have been implicated in the decline of some species, and causes of apparent declines in others remain enigmatic. Little data exist on the status of snake populations in tropical regions of the world; however, as with other taxa, snakes are undoubtedly suffering as a result of rampant habitat destruction occurring in tropical regions. A relatively novel, but poorly understood threat to snakes in these regions is the phenomenon of cascading declines discussed previously. In Asia, snakes face greater pressure from exploitative use than in other regions of the world, with estimated millions of snakes harvested annually from China and other regions of Southeast Asia (Zhou and Jiang 2004, 2005; Brooks et al. 2007). Finally, the myriad introduced exotic species in Australia (e.g., foxes, feral cats, cane toads) pose serious threats to the fragile ecosystems of that island continent. Indeed, as noted previously, cane toads alone have been suggested to threaten as much as 30% of Australia’s terrestrial snake species (Phillips et al. 2003). Although our knowledge of the status of global snake populations remains woefully inadequate, increasing awareness of the importance of snakes as top predators in many ecosystems (e.g., Ineich 2007) and advances in methodology for studying snake populations (Dorcas and Willson in press) will undoubtedly increase our ability to effectively conserve snake populations in future decades.

3.4  Conclusion Continuing to determine the status, distribution, and basic ecology of many reptiles is of paramount importance. Although low detectability of reptiles, and subsequently poor awareness of declines in their populations, may hamper research and conservation efforts, ongoing advances in field methodology, mark-recapture analyses, and our understanding of reptile life histories and behaviors should

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continue to improve our knowledge of the status and distribution of many reptiles. As recognition of declining reptile populations increases, determining causes of those declines should also become a primary goal. Some populations and species may be affected by one or a few factors, but multiple interacting stressors or threats likely affect many reptile populations. Numerous studies have demonstrated direct effects that environmental contaminants have on reptiles. But the many ways that environmental contaminants could exacerbate ongoing declines from other threats, such as disease and parasitism, habitat loss, and introduced invasive species, remain underappreciated. The hope that reptile declines will tail off or that highly imperiled species will be able to claw their way back from near extinction rests on a full understanding of the plight of reptiles, threats to their populations, causes for their declines, and effective mobilization of conservation resources.

Acknowledgments Manuscript preparation was aided by the Environmental Remediation Sciences Division of the Office of Biological and Environmental Research, US Department of Energy, through Financial Assistance Award DE-FC09-96SR18546 to the University of Georgia Research Foundation.

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4

Ecotoxicology of Amphibians and Reptiles in a Nutshell Greg Linder, Christine M. Lehman, and Joseph R. Bidwell

Contents 4.1 Historic Developments: Amphibians and Reptiles in Ecotoxicology..................................... 70 4.1.1 Amphibians.................................................................................................................. 70 4.1.2 Reptiles........................................................................................................................ 70 4.2 Contaminant Exposure Pathways for Amphibians and Reptiles............................................. 71 4.3 Toxicity Assessment: Laboratory Studies............................................................................... 72 4.3.1 Toxicity Tests Using Amphibians................................................................................ 72 4.3.2 Toxicity Tests Using Reptiles......................................................................................80 4.4 Toxicity Assessment: Mesocosm Studies................................................................................ 81 4.5 Toxicity Assessment: Field Studies.........................................................................................84 4.6 Biomarkers of Exposure and Effect in Amphibians and Reptiles........................................... 86 4.7 Recurring and Emerging Issues: Future Challenges for Toxicologists Studying Amphibians and Reptiles......................................................................................................... 89 4.7.1 Recurring Issues in the Ecotoxicology of Amphibians and Reptiles.......................... 89 4.7.2 Emerging Chemical Contaminants.............................................................................90 4.7.3 Ecotoxicology and Multiple Stressors......................................................................... 91 4.8 Summary.................................................................................................................................92 Acknowledgments............................................................................................................................. 93 References......................................................................................................................................... 93 Concern over the localized reduction of amphibian populations and the potential role that chemical stressors were playing in these declines was expressed nearly 30 years ago (Gibbs et al. 1971; Birge et al. 1980; Bury 1999), an observation recently revisited for reptiles (Gibbons et al. 2000). However, it has only been over the last decade that amphibians, and to a lesser extent reptiles, have been recognized as vertebrates truly unique from other terrestrial wildlife and the fishes. The burgeoning interest in the ecotoxicology of the herpetofauna has been driven in part by their continued population declines at a global scale and a growing appreciation that terrestrial wildlife such as birds and mammals and aquatic vertebrates such as fish may not adequately represent contaminant exposure experienced by amphibians and reptiles during the various phases of their life cycles. For most amphibians, this includes laying permeable eggs in water, an aquatic larval stage, a physiologically demanding period of metamorphosis from larva to juvenile, and an adult stage that occurs in both aquatic and terrestrial habitats. Even more recently, the importance of reptiles as ecological receptors in both aquatic and terrestrial habitats has been recognized. While lacking the distinctly bimodal life cycle seen in most amphibians, the long lives, philopatric tendencies, and for some species, amphibious lifestyles exhibited by reptiles may expose them to a variety of contaminants over very long time periods (Sparling et al. 2000a). The continuing increase in ecotoxicological research activities with herpetofauna has supported synthesis publications such as Sparling et al. (2000b), Linder et al. (2003a, 2003b), Campbell and Campbell (2000, 2001), and Gardner and 69

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Oberdörster (2006), and compilations of ecotoxicological data (Devillers and Exbrayant 1992; Pauli et al. 2000). In this chapter, we provide a snapshot of the existing literature pertaining to the ecotoxicology of amphibians and reptiles, particularly for species reliant to varying extents on aquatic environments. We will briefly consider the historic context of these organisms as players in the discipline of ecotoxicology. Then, we will discuss their current roles in guiding ecotoxicologists toward a future where these previously undervalued “orphan groups” are more fully appreciated as critical components of a wide range of aquatic, wetland, and terrestrial habitats.

4.1 Historic Developments: Amphibians and Reptiles in Ecotoxicology 4.1.1 Amphibians The eggs and larvae of anuran amphibians have a long history of use as models for the study of early vertebrate development, including the effects of materials that disrupt developmental patterns (e.g., Chang et al. 1954; also see Callery 2006 for review). In the 1970s and early 1980s, an increasing number of studies began evaluating the effects of organic and inorganic chemicals on fish and invertebrates (e.g., Mount and Brungs 1967; Cairns and Dickson 1973; Committee on Methods for Acute Toxicity Tests with Aquatic Organisms 1975; Birge and Black 1977; Birge 1978; Peltier 1978) and amphibians (e.g., Sanders 1970; Cooke 1972; Johnson 1976; Birge et al. 1980) for the sake of determining the risk these compounds might pose to organisms in the field. ASTM standards (e.g., E-729, Standard Practice for Conducting Acute Toxicity Tests with Fishes, Macroinvertebrates, and Amphibians) initially published in 1980 were instrumental in setting the stage for developing tools for toxicity assessment. Also, as reviewed by Birge et al. (2000), some of this work ultimately led to methods that were adopted by the USEPA to support biomonitoring programs promulgated as part of the Clean Water Act (although amphibians were not ultimately included as species commonly used in these tests). At this same time, Dumont and colleagues (Dumont et al. 1979, 1983; Bantle 1995) were undertaking work with Xenopus laevis to develop screening tests that could be used to indicate effects of process waters and complex mixtures associated with oil and gas extraction. These methods development efforts ultimately lead to the protocol for the Frog Embryo Teratogenesis Assay–Xenopus (FETAX). Beyond these simple laboratory-based tools used to evaluate chemical effects on amphibians, Semlitsch and Bridges (2005) have advocated incorporating realism into experiments (e.g., utilizing native species in toxicity tests), focusing on differences in life modes and rates of development (e.g., direct vs. indirect developers), incorporating greater genetic variation of test organisms into studies, increasing the spatial scale of studies, and examining direct vs. indirect effects of contaminants. They also encourage exploring the biological links between ecotoxicological studies and conservation, links that are potentially disrupted when challenged, yielding adversely affected regulation of species populations and community structure. Contemporary studies are steering away from single-species, single-contaminant approaches that dominated early investigations focused on the ecotoxicology of herpetofauna, and go beyond observations that populations of some amphibians were highly reliant on aquatic habitats for their early developmental stages. While chemicals entering aquatic habitats remain stressors critical to long-term sustainability of amphibian populations, opting to a multiple stressors approach for managing amphibian populations may be more beneficial to their long-term sustainability.

4.1.2 Reptiles In contrast to the nearly 40-year history of amphibians being used in aquatic toxicology, the role of reptiles has unfortunately been relatively minor and more episodic (Hopkins 2000), despite the

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fact that reptiles have been undergoing population declines much the same as amphibians (Gibbons et al. 2000). Compilations of toxicity data are available (see Sparling et al. 2000b; Pauli et al. 2000), but relatively underdeveloped, given the sparse literature to support such reviews. Until recently, the paucity of toxicity data stemmed from the lack of standard laboratory test procedures for evaluating toxic effects of chemicals on this taxon. While more standard methods for conducting toxicity tests with a reptile model have recently become available (Brasfield et al. 2004), the focus is most relevant to terrestrial habitats. While studies with amphibians have consistently outpaced those with reptiles, especially with respect to characterizing toxicity data, the evaluation of bioaccumulation appears better developed for reptiles than for amphibians. As noted in Sparling et al. (2000a), published works focused on bioaccumulation of chemicals demonstrate the significant role that monitoring of reptiles such as turtles has played in the evaluation of aquatic contaminants. For example, earlyexposure assessments using wildlife frequently documented tissue residues in field-collected turtles (e.g., Bishop et al. 1998; Golet and Haines 2001), and studies of alligators (Alligator mississippiensis) in Florida were important for characterizing uptake and accumulation of mercury (e.g., Heaton-Jones et al. 1997) and endocrine-disrupting chemicals (EDCs; e.g., Guillette et al. 1994, 1995). While chemical effects and exposure data for herpetofauna will undoubtedly continue to increase in the future, the paucity of data underscores data gaps clearly apparent for reptiles. Work focused on amphibians should also increase, or at the very least, continue at its current pace.

4.2 Contaminant Exposure Pathways for Amphibians and Reptiles The complex life cycle of most amphibians leads to diverse ways in which they are exposed to environmental contaminants. As discussed by Birge et al. (2000), uptake of chemicals can begin shortly after egg deposition, as water moves into the egg capsule. There is actually some indication that the eggs themselves may receive a residue load from maternal transfer (Birge et al. 2000; Kadokami et al. 2002, 2004; Hopkins et al. 2006), a phenomenon reported for reptiles (see next paragraph), but much less so for amphibians. As larvae and tadpoles, uptake of waterborne chemicals across the permeable skin is an important source of exposure. The skin continues to play a role in chemical uptake by most adult amphibians as a result of its continued function as a respiratory surface, with the lungs also providing possible routes for volatile compounds. Uptake of chemicals through food may be important throughout the life cycle, although the importance of this route may vary. For example, within their lifetime, many anurans exist on 2 trophic levels, eating algae as tadpoles and invertebrates as adults. Adult amphibians can burrow into sediment and soils during hibernation and aestivation, making uptake of contaminants from these solid phases (across their permeable skin) a potential avenue of exposure as well (see Chapter 5, this volume; Boutilier et al. 1992; Larsen 1992; Shoemaker et al. 1992). As noted earlier, maternal transfer of both organic and inorganic contaminants has been described for reptiles, with the eggs often used to indicate maternal contaminant burdens (Pagano et al. 1999; Nagle et al. 2001). This may lead to significant impacts on early life stages as demonstrated by Rauschenberger et al. (2004), who found that maternal transfer of organochlorine pesticides was associated with reduced egg and embryo viability in American alligators. Although not singly dependent on aquatic habitats to complete early developmental stages of their life cycle, many reptiles nest in terrestrial and wetland environments in close association with surface waters, and the eggs, when buried in soils, may accumulate chemicals from the surrounding matrix in association with water that is imbibed during the course of development. Moeller (2004) reported significant accumulation of lead, cadmium, and zinc in embryos of red-eared slider turtles (Trachemys scripta) derived from eggs that had been incubated on contaminated substrates. Since the skin of most reptiles is relatively impermeable, dermal uptake of chemicals may not be a particularly important route of exposure for these organisms. However, some aquatic turtles rely on water held in their

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buccal cavity for oxygen uptake (Withers 1992), and this may also provide a pathway for entry of dissolved chemicals. Whereas terrestrial exposure may seem less important for some herpetofauna, McDiarmid and Mitchell (2000) comment on the relatively long distances that species of both amphibians and reptiles may move daily, seasonally, and annually, a feature that can significantly influence the potential to be exposed to contaminants outside the aquatic environment. In adults, bio­ accumulation may be dominated by dietary exposures, yet chemicals of the exposure matrix (e.g., metals) may be readily absorbed across dermal epithelia or volatilize from a solid or liquid phase into air, providing another means of exposure (see, for example, Noble 1931; Boutilier et al. 1992; Shoemaker et al. 1992; Duellman and Trueb 1994; Stebbins and Cohen 1995 regarding cutaneous respiratory surfaces and potential routes of metal uptake). Many variables affect the magnitude of bioaccumulation in terrestrial exposures to adults or early developmental stages (e.g., in reptile eggs), and the transfer of chemicals within food chains may be conveniently described by transfer coefficients or functions that characterize the relationships among trophic levels (Pastorok et al. 1996; Pascoe et al. 1996; Linder et al. 1998; Linder and Joermann 1999). These factors may be abiotic, as reflected by physicochemical characteristics of a chemical and an exposure matrix (sediment or soil), or biological in character, as captured by life-historydependent attributes related to gastrointestinal or nutritional physiology, foraging, or feed preference (see Young 1981; Larsen 1992; Hamelink et al. 1994; Langston and Spence 1995; Linder et al. 2002).

4.3  Toxicity Assessment: Laboratory Studies 4.3.1 Toxicity Tests Using Amphibians A key focus of applied toxicology is to assess the risk that potential chemical stressors may pose to natural populations. Laboratory studies may be used in an a priori sense to help predict the effects of a chemical before it is released into the environment, or to reassess the risk of materials as additional data regarding their environmental effects become available. For example, basic laboratory tests with a common herbicide formulation indicated the potential for adverse effects on tadpole stages in Australian anurans and resulted in revised labeling guidelines and restrictions on its use (Mann et al. 2003). Laboratory toxicity tests are also used extensively for regulatory assessments of wastewater discharges, and Cooney (1995) reviewed some of the basic methods, test species, and statistical endpoints relevant to freshwater tests. The list of response parameters that can be evaluated in a laboratory setting is extensive, and those used in a particular study are obviously influenced by study objectives. Mortality and growth of test organisms are common endpoints that may be used with most test species (ASTM 2005a, 2005b), while endpoints more specific to amphibians include the length of the larval period and size at metamorphosis. Examples of laboratory tests using native North American amphibian species and associated endpoints are included in Table 4.1. As further described later. FETAX can be used to evaluate the incidence of malformation in addition to growth and mortality (ASTM 2005c). Other endpoints that have been used in amphibian tests include measures of performance and/or behavior after exposure to chemical stressors, such as swim speed, the ability to avoid predators, general activity levels, and time spent feeding (Bridges 1997; Savage et al. 2002; Richards and Kendall 2003; Broomhall 2004; Widder and Bidwell 2006, 2008). Recent laboratory work using high-speed video to investigate the subtleties of the escape response (C-start response) in fish and aquatic amphibian stages (e.g., Azizi and Landberg 2002) should also prove useful for investigating sublethal contaminant effects on escape responses. Finally, various biochemical and physiological “biomarkers” can and have been investigated in the laboratory setting, as discussed further later. Laboratory toxicity tests using amphibians have been conducted with embryo and larval stages (collectively considered early life stage [ELS] tests), metamorphs, juveniles, and sexually mature

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Atrazine Iridovirus

Multiple agricultural chemicals

Atrazine Nitrate

Atrazine Density Glyphosate

Fungicide

Atrazine

Cadmium Malathion Predation Carbaryl Predation

Laboratory

Laboratory

Laboratory

Laboratory

Laboratory

Laboratory

Laboratory Laboratory

Stressor

Laboratory

Study Type

• • • • • • • • • • •

• • • • • • • • • • • MMM, LLP Activity Survival Abnormalities % metamorph Behavior MMM, LLP MMM, LLP Gonad development Hibernation success Acute survival

Development Growth Survival Development Growth Gonadal development Immune function MMM, LLP Gonadal differentiation Sex ratio Postexposure survival

Endpoint

Juveniles Tadpoles

Embryo to metamorph

Tadpoles to metamorph

Embryos Larvae Tadpoles to metamorph

Tadpole to metamorph

Tadpole to metamorph

Larvae

Life Stage

Bufo americanus Bufo americanus Hyla versicolor Rana catesbeiana R. clamitans R. pipiens R. sylvatica

Rana clamitans

Rana temporaria

Rana cascadae

Ambystoma barbouri

Rana pipiens

Rana pipiens Xenopus laevis

Ambystoma macrodactylum

Species

Table 4.1 Examples of Recent Laboratory, Mesocosm, and Field Studies Using Native North American Amphibian Species

James et al. 2004b Relyea 2004 Relyea 2003

Coady et al. 2004

(continued)

Teplitsky et al. 2005

Cauble and Wagner 2005

Rohr et al. 2006

Orton et al. 2006

Hayes et al. 2006

Forson and Storfer 2006

Reference

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Atrazine Endosulfan Competition Predation Carbaryl Competition Predation Copper Pathogen Copper Pathogen

Mesocosm

Mesocosm

Mesocosm

Mesocosm

Mesocosm

Atrazine Limited food Hydroperiod

Roundup®

Mesocosm

Mesocosm

Mesocosm

Mesocosm

Endosulfan Temperature Carbaryl Nitrates Cadmium Tissue residue Carbaryl Competition

Stressor

Laboratory

Study Type Predator avoidance Survival MMM, LLP Survival MMM, LLP Survival MMM, LLP Lipid reserves % metamorphs Survival

• • • • • • • •

MMM, LLP Survival Fluctuating asymmetry MMM, LLP Survival Behavior MMM, LLP Survival

• MMM, LLP • Survival

• MMM, LLP • Activity • Predator avoidance

• • • • • • • • • •

Endpoint

Egg to metamorph

Tadpole to metamorph

Tadpole to metamorph

Tadpole to metamorph

Tadpole to metamorph

Ambystoma barbouri

Bufo fowleri Hyla chrysoscelis

Hyla chrysoscelis

Rana sphenocephala

Bufo americanus Hyla versicolor Pseudacris crucifer Rana pipiens Rana sylvatica Rana sylvatica

Larvae Metamorphs

Larva to metamorph

Bufo americanus Rana sphenocephala Ambystoma maculatum A. opacum

Rana clamitans

Litoria citropa

Species

Tadpole to metamorph

Tadpoles to metamorphs

Tadpoles

Life Stage

Table 4.1 (Continued) Examples of Recent Laboratory, Mesocosm, and Field Studies Using Native North American Amphibian Species

Rohr et al. 2004

Parris and Cornelius 2004

Parris and Baud 2004

Mills and Semlitsch 2004

Rohr and Crumrine 2005

Relyea 2005

Metts et al. 2005

James et al. 2005

Boone et al. 2005

Broomhall 2002

Reference

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Carbaryl

Atrazine Carbaryl Competition Hydroperiod UV radiation Carbaryl Atrazine Carbaryl Endosulfan Octylphenol Limited food Carbaryl Competition Predation Hydroperiod

Mesocosm

Mesocosm

Carbaryl

Atrazine

Atrazine

Retinoids

OC pesticides

Mesocosm

Field survey

Field survey

Field survey

Field survey

Mesocosm

Mesocosm

Mesocosm

Mercury

Mesocosm

MMM, LLP Survival Activity Growth rate Hatching Survival

MMM, LLP Malformity Survival MMM, LLP Survival MMM, LLP Survival

MMM, LLP Survival Aromatase activity Hormones

• Weight • Tissue residues • Tissue residues

• Gonad histology

• • • •

• MMM, LLP • Survival

• • • • • •

• • • • • • •

Tadpoles

Frogs

Juveniles Adults

Frogs

Tadpoles to metamorphs

Tadpoles to metamorphs

Egg to metamorph

Tadpoles to metamorphs

Tadpoles to metamorphs

Tadpoles to metamorphs

Tadpole to metamorph

Rana lessonae R. esculenta

R. catesbieana

Rana catesbieana R. clamitans R. pipiens Rana catesbieana R. clamitans R. pipiens

Rana blairi R. clamitansi R. sphenocephala B. woodhousei Notophthalmus viridescens Rana clamitans

Ambystoma barbouri

Ambystoma maculatum A. texanum Bufo americanus Rana sphenocephala Rana sphenocephala

Rana clamitans

Rana sphenocephala

Fagotti et al. 2005

Berube et al. 2005

(continued)

Murphy et al. 2006b

Murphy et al. 2006a

Boone et al. 2001

Boone and Semlitsch 2002

Rohr et al. 2003

Bridges and Boone 2003

Boone and James 2003

Boone and Bridges 2003

Unrine et al. 2004

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Environmental exposure

Environmental exposure

PCBs Toxaphene Environmental exposure OCs, PCBs Metals Environmental exposure

Release®

Environmental exposure

Glyphosate

UV Nitrates

Field survey

Field survey

Field survey

Field manipulation

Field manipulation

Field manipulation

Field manipulation

Field manipulation

Field survey Field survey

Environmental exposure

Field survey

Stressor

Environmental exposure Coal combustion waste

Field survey Field survey

Study Type

• • • • • •

• • • • • • • • • • • •

• • • • • • • Population persistence Residues Deformities Hatching Growth Species richness Survival Tissue residues Avoidance Growth Survival Cytochrome p450 activity Avoidance response Growth Survival Mass Length Survival

Residues Population status Species richness Reproductive success Species richness Tissue PCB residues Tissue residues

• Tissue PCB residues • Tissue residues

Endpoint

Species

Tadpoles Larvae

Tadpoles

Ambystoma macrodactylum Hyla regilla

Rana clamitans R. pipiens

Rana clamitans melanota

Rana clamitans R. pipiens

Tadpoles

Tadpoles

Rana clamitans R. pipiens

Various species Rana clamitans

Hyla regilla

Various species

Various species

Rana muscosa

Rana temporaria Bufo terrestris Rana sphenocephala

Eggs Embryos Tadpoles

Frogs Frogs

Tadpoles

Frogs

Frogs

Tadpoles Tadpoles Metamorphs Frogs Frogs

Life Stage

Table 4.1 (Continued) Examples of Recent Laboratory, Mesocosm, and Field Studies Using Native North American Amphibian Species Reference

Hatch and Blaustein 2003

Wojtaszek et al. 2004

Jung et al. 2004

Wojtaszek et al. 2005

Karasov et al. 2005

Davidson et al. 2002 Gillilland et al. 2001

Angerman et al. 2002

DeGarady and Halbrook 2003

Knutson et al. 2004

Fellers et al. 2004

Hofer et al. 2005 Roe et al. 2005

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Environmental extracts

Environmental extracts

Agricultural chemicals

Atrazine Environmental and in-lab exposure

PCBs

Agricultural runoff Parasite Carbaryl

PCB-laden sediment

Integrated study

Integrated study

Integrated study

Integrated study Integrated study

Integrated study

Integrated study

Integrated study

Integrated study

Survival Species richness Tissue residue Immune function Deformities MMM, LLP Survival

• Behavior • Survival

• • • • • • • Tadpole Metamorph Eggs Hatchlings Tadpoles Tadpoles

Eggs to frogs

• Deformity • Survival • MMM, LLP

Frogs

Tadpoles to metamorphs

Tadpoles to metamorphs

Eggs Tadpoles

Frogs Eggs to tadpoles

Abnormalities Behavior Hatching Tissue residue MMM, LLP Survival Abnormalities MMM, LLP Survival Abnormalities Population status

• Gonadal morphology • Hatching

• • • • • • • • • • •

Rana sylvatica

Hyla versicolor

Rana sylvatica

R. utricularia

Rana pipiens

Rana aurora

Bufo canorus Rana aurora draytonii R. boylii R. cascadae R. muscosa Rana pipiens Ambystoma gracile

Hyla arenicolor Pseudacris crucifer P. regilla Rana pipiens

Gastrophryne carolinensis

Savage et al. 2002

Saura-Mas et al. 2002

Kiesecker 2002

Glennemeier and Begnoche 2002

Hayes et al. 2003 de Solla et al. 2002

Davidson 2004

Bridges et al. 2004

Bridges and Little 2005

Hopkins et al. 2006

Note: FETAX studies are not included, although the number of publications using the protocol is extensive. See text for further discussion of FETAX. MMM = mass at metamorphosis, LLP = length of the larval period.

Coal combustion waste

Integrated study

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adults. Of these stages, embryos and larvae have been used most commonly to evaluate the effects of aquatic contaminants. Part of the reason for this is the common assumption in aquatic toxicology regarding greater sensitivity of early life stages, and there are a number of studies that have indicated that tadpoles are more sensitive to contaminants than eggs and postmetamorphic adults (Ralph and Petras 1998; Mann and Bidwell 1999). However, studies that compared the sensitivity of different life stages of amphibians with a finer degree of resolution have indicated enhanced effects during metamorphosis, which may be due to the physiological demands associated with the “reorganization” that organisms experience through developmental time (Howe et al. 1998; Natale et al. 2000; Fort et al. 2004a). Another factor leading to the widespread use of aquatic stages (in particular, tadpoles) is their greater availability (particularly in the case of field-collected organisms) and use for waterborne exposures. While an early approach by Birge and Black (1977) placed fish and amphibian embryos in contact with contaminated sediments, the majority of laboratory tests with amphibian larvae have focused on exposure via the water column (for example, see the methods described in ASTM E-729 [2005a]). Since postmetamorphs and adults of many species drown if held in water constantly, a key challenge in evaluating contaminant effects in these stages is the development of an appropriate exposure system. Some particularly novel work has considered uptake through the diet and dermal exposure in active and hibernating adults (James 2003; James et al. 2004a, 2004b). Bantle (1995) discussed the importance of standardization of test methods, since it facilitates comparison of species sensitivity to the same contaminant. In this regard, FETAX and its derivatives are probably the most popular of the laboratory test methods that have used amphibians to assess single chemicals and complex mixtures. Bantle (1995), Dumont et al. (2003), and Fort et al. (2003) have summarized the development of FETAX, which became available as a standard aquatic toxicity test as ASTM E-1439 in 1991 (ASTM 2005c). Interestingly, the method was originally designed in the 1970s and early 1980s as an outcome of methods development intended to screen chemicals that were potential human developmental hazards (Bantle 1995). FETAX, however, exceeded these intentions and contributed to heightening focus on amphibians as receptors incorporated into the ecological risk assessment. As a test species, X. laevis and the smaller X. tropicalis present life history attributes amenable to laboratory testing, since they can be easily maintained in the laboratory, can be hormonally induced to reproduce throughout the year (unlike North American amphibians, which are seasonally reproductive and are usually not as amenable to laboratory culture), and can provide large numbers of embryos for testing. Exposures are initiated at early blastula and continue through primary organogenesis to ensure a baseline assessment of early life stage effects. Endpoints include mortality (which is evaluated over the course of the exposure), embryo growth (as length and mass), number of abnormal embryos, and type of abnormailities observed at the end of the test. These data provide LC50 and EC50 data (median lethal concentration and median effective concentration for terata based on number of abnormal embryos among the survivors, respectively). X tropicalis completes its life cycle in a shorter amount of time than X. laevis, which has stimulated interest in its use in life cycle toxicity tests (Fort et al. 2004b). These tests can be important for understanding the reproductive effects that contaminants have on amphibians. For example, Lienesch et al. (2000) observed adverse effects of sublethal cadmium exposures in female X. laevis oocytes at all stages of oogenesis, including greatly increased numbers of atretic oocytes and other indicators that oocytes would be significantly reduced in their viability. Christensen et al. (2004) describe an amphibian sperm inhibition toxicological test (ASITT) that examines the effects of contaminants on Xenopus sperm motility. Other genera besides Xenopus may also be tested with the FETAX protocol. For example, Rana catesbiena, R. pipiens, Bufo fowleri, and B. americanus are alternative test organisms identified in ASTM E-1439. When applying the protocol to other species, recall that development is highly temperature dependent and varies greatly among species, so if developmental stage serves as the endpoint for test termination (e.g., for FETAX, Stage 46 as characterized by Nieuwkoop and Faber [1956]), exposure times may vary among taxa.

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Another test system that utilizes early amphibian life stages in a fashion similar to FETAX is the AMPHITOX method described by Herkovits and Pérez-Coll (2003). This procedure actually comprises a group of different types of bioassays that evaluate acute (AMPHIACUT), short-term chronic (AMPHISHORT), and chronic (AMPIRCHRO) exposures. By plotting toxicity endpoints derived from these tests, a family of toxicity curves can be generated from exposure data collected for 24 hours to 14 days. Conventional measures of toxicity, such as no observed effect concentrations (NOECs), lowest observed effect concentrations (LOECs), and median effect concentrations (LC50s), may also be calculated. By employing an early life stage test (AMPHIEMB), developmental effects may also be evaluated. Results of AMPHITOX suggest that the toxicity of a wide range of environmental samples may be evaluated by selecting the most appropriate toxicity test included in the test suite, with the test selection determined in part by an initial screening level evaluation of toxicity of the sample and the endpoint of concern. The role of laboratory toxicity studies with amphibians can and should extend beyond the assessment of waterborne stressors. For example, while sediment studies are relatively uncommon in amphibian ecotoxicology, tadpoles of many species spend much of their time in the sediment and even forage on decomposing organic matter found there. Thus, contaminants that are bound to sediment (and absent or in low concentrations in the water column) can serve as an important source of exposure via both dermal and dietary routes of exposure. For example, Lehman et al. (personal communication) found that tadpoles reared with sediment from contaminated sites in the Alaskan Kenai Peninsula were smaller upon metamorphosis than tadpoles reared under control conditions. Furthermore, because amphibians have permeable skin, contact with sediments may represent an important route of uptake via dermal exposure. This is the case with developing tadpoles as well as hibernating adults and juveniles that burrow into sediments. Bleiler et al. (2004) have published methods that address the current deficit in sediment testing procedures using amphibians. Studies incorporating sediment exposures should be designed to allow comparisons of toxicity similar to those of invertebrate test methods for assessing contaminated sediments (ASTM 2005d). Standard laboratory studies following the FETAX protocol have been used to evaluate extracts from soils and sediments collected from contaminated areas (Fort et al. 2001). Tests have also been conducted to test soils, soil eluates, and sediments directly by suspending embryos over a sediment sample in an exposure chamber (Birge and Black 1977; Fort et al. 1999). Outcomes of these tests indicated that toxic constituents may not be tightly bound to the solid-phase material. This type of work can play an important role in evaluating the presence of an existing contaminant load in wetland habitats and help direct management or remedial decisions (e.g., Hutchins et al. 1998). Some species of amphibians are not amenable to laboratory testing due to factors such as low survival in a laboratory setting, difficulty in capturing sufficient numbers of individuals to conduct statistically defensible tests, or in the case of threatened or endangered populations, legal and ethical considerations associated with continued depletion of natural populations. As such, surrogate test organisms are often used to generate response data, with Xenopus often the species of choice for many of those wishing to examine the effects of contaminants on amphibians. The key advantage of these organisms is that they are easily kept in the laboratory and can yield eggs for tests year-round. However, Birge et al. (2000) stated that the higher tolerance of Xenopus to chemical stressors, as compared to other amphibian species, makes it less suitable for use in routine testing for aquatic risk assessment. In their evaluation of FETAX for use in ecological risk assessments, Hoke and Ankley (2005) concluded that risk assessments using acute hazard data with traditional laboratory species were more protective of native amphibians than assessments based on hazard data from FETAX. On the basis of the available comparative toxicity data, X. laevis sensitivity appears to lie mid-range, tending toward being less sensitive than many North American species tested (Birge et al. 2000). These observations inevitably vary as a function of toxicant and endpoint. For example, Mann and Bidwell (2000) found that Xenopus was the most sensitive of the organisms they evaluated in FETAX and modified FETAX assays of an agricultural surfactant, and Hoke and

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Ankley (2005) observed that growth was a more sensitive endpoint in the assay than malformation or survival. Although the species has invaded areas of Florida and southern California, Xenopus is not native to North America and may be limited in use as a surrogate for North American anurans. While native species are often only seasonally available, the ecological relevance of native forms may outweigh this issue, particularly in cases of site-specific risk assessments. A key objective in this regard is to extend the existing comparative toxicity data for native species, since the use of surrogates (native forms) is likely to remain an important approach. One of the most significant criticisms of laboratory toxicity tests is their potential to inadequately predict the effect chemical stressors might have on organisms in the field (Burkhart et al. 2003). Certainly, the “real world” situation imposes a suite of stressors that the often simplistic exposure environment of the laboratory fails to address. Still, laboratory toxicity tests continue to play an important role in evaluating chemical effects. A significant advantage of these procedures is that they allow control of potentially confounding physical (e.g., temperature) and chemical (e.g., pH and hardness when using formulated diluents) variables, making them critical for establishing causeand-effect relationships between the presence of contaminants and the response of organisms. Laboratory toxicity tests have also played a valuable role in characterizing the sensitivity of different amphibian groups to chemical stressors (Bridges and Semlitsch 2005). Although standardization of test methods is important in this regard, within the scope of amphibian declines most effects on populations may not be attributable to lethal concentrations of contaminants present in a habitat; hence, some laboratory-generated toxicity estimates (e.g., median lethal concentrations) may be relatively limited in applications far removed from evaluation of field settings. However, examining alternative endpoints related to adverse effects on life history traits may be more applicable to characterizing ecologically acceptable levels of a contaminant rather than relying only on traditional lethality-based endpoints. For instance, the length of the larval period and mass at metamorphosis are traits critical in determining an individual’s fitness (i.e., survival and future reproductive success). Amphibians that metamorphose at larger sizes and earlier in the season have a greater chance of surviving over winter and will reproduce at younger ages (Smith 1987; Semlitsch et al. 1988). Further, a short larval period is especially important to amphibian species breeding in temporary ponds, where any factor that lengthens the larval period, such as the presence of an environmental contaminant, can lead to mortality due to desiccation or prolonged exposure to predators. Many contaminants alter life history traits at concentrations well below those that express themselves by decreased survival. The use of novel exposure scenarios in the laboratory has also indicated the complexity of the organismal response to chemicals, as demonstrated by Relyea and Mills (2001), who found the presence of a predator enhanced the toxicity of a pesticide to gray treefrog (Hyla versicolor) tadpoles.

4.3.2 Toxicity Tests Using Reptiles With regard to ecotoxicology studies focused on herpetofauna, amphibians have outpaced reptiles, although the latter have accounted for an increased number of citations over the past 10 to 15 years. As part of their effort to initially characterize the state of the science of ecotoxicology for amphibians and reptiles, Sparling et al. (2000a) completed a literature search that spanned more than 25 years between 1972 and 1998, and found that only 1% to 2% of the publications focused on vertebrates concerned reptiles and their role in evaluating or monitoring chemicals in the environment. As shown in Chapter 1 of this book, this relative standing of reptiles in ecotoxicological research has not changed appreciably. This is likely attributable to amphibians, in general, being more amenable to toxicity tests than reptiles, especially during the formative years of aquatic toxicology. Many anurans lay thousands of eggs at a time, providing ample numbers of tadpoles (i.e., test animals and adequate replication) for experiments. Urodeles commonly have lower reproductive output than anurans, but can still produce offspring in sufficient numbers for aquatic toxicity

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tests. Reptiles, by comparison, are more difficult to collect and produce relatively fewer offspring per clutch than most amphibian species. At present, toxicity test procedures for reptiles are largely focused on snakes and lizards (see Table 4.2 for examples). Toxicity testing with selected terrestrial species of lizard (primarily eastern and western fence lizards, Scaphiopus undulatus and S. occidentalis, respectively) is increasingly reported in the literature, yet requires standardization for its widespread application as a tool to inform risk management. No test systems are available to evaluate reptiles that are predominantly aquatic species. In general, slow growth rates, long life cycles, and complex relationships between size and age at sexual maturity are some of the challenges in establishing captive breeding populations of aquatic and semiaquatic reptiles for developing standard test methods comparable to those available for fence lizards or the common aquatic vertebrate test species. Although Crews et al. (2003) have suggested that red-eared sliders (Trachemys scripta elegans) be applied in screening tests for endocrine-disrupting chemicals, no standardized test has been developed. Turtle eggs, however, have also been used to study maternal transfer of contaminants, uptake of contaminants from nest substrate, and effects of those contaminants on hatchlings (Table 4.2). From a practical perspective, tests using fence lizard eggs may be preferred over those of turtles, since fence lizards will grow and reproduce in the laboratory, and their eggs would be more readily available. In contrast, turtle eggs are usually only seasonally available from field collections, where they may have been exposed to chemicals while lying in the nest or maternally transferred. For studies focused on endpoints related to endocrine-disrupting chemicals, external differences in the sex of fence lizards are more easily determined early in the animal’s development; in turtles, dissection is usually required to determine sex, particularly in juveniles. Fence lizards also appear to be good candidates for use in a reptilian reproductive toxicity test because populations mature quickly and reproduce well in captivity. Within ecological contexts, the herpetofauna include many species that potentially link aquatic habitats and immediately adjacent wetland or terrestrial habitats. For example, semiaquatic reptiles, in particular their early life stages developing in ovo, have been used to evaluate chemicals in seasonally hydric soils. Although saturated soils may not be appropriate for egg incubation, within ecological contexts toxicity evaluations framed within the life history patterns of semiaquatic reptiles better serve the ecological risk assessment process than alternatives simply focused on chemical analysis of substrates that yield modeled extrapolations based on other terrestrial vertebrates. Despite advances since Sparling et al. (2000b) initially summarized the state of the science for the ecotoxicology of amphibians and reptiles, development of comparative toxicity data for reptile orders continues to lag and remains one of the long-term objectives identified by Hopkins (2000) to advance reptilian toxicology.

4.4  Toxicity Assessment: Mesocosm Studies Interpretation of laboratory data within the context of ecological risk requires validation and confirmation of laboratory toxicity results under field conditions, with the goal being to minimize laboratory-to-field extrapolation errors. Effects of many chemical contaminants may be altered by field conditions that vary with habitat, making it important to evaluate toxicity while accounting for environmental factors potentially influencing exposure or effects. For example, UV radiation can cause some environmental chemicals to be more toxic and, in other instances, it may promote degradation of contaminants to less toxic forms (Bridges et al. 2004). In aquatic habitats, hydroperiod and larval density may influence the effects of chemical contaminants on amphibian larvae (e.g., carbaryl; Boone and James 2003) by lengthening the larval period and causing animals to be smaller at metamorphosis. Predation may also influence toxic effects linked to chemical exposure, potentially yielding increased toxicity for some chemicals under field conditions (see, for example, Relyea and Mills 2001 as briefly noted later).

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Carbaryl

Toxaphene

Trace elements

Metals Metals Metals

Metals

OCs Mercury Selenium OCs Environmental exposure

Laboratory

Laboratory

Laboratory

Laboratory Field survey Field survey

Field survey

Field survey

Field survey Field survey Field survey

Carbaryl

Stressor

Laboratory

Study Type

Tissue residues Tissue residues Tissue residues

Tissue residues

Tissue residues

Morphological Physiological Tissue residues Behavior Growth Body condition Egg residues, hatchling mass Egg residues Tissue residues

Swimming performance

Swimming performance

Endpoint

Adults Eggs Eggs

Adults

Adults

Eggs Eggs Adults

Eggs Hatchlings Juveniles

Neonate

Neonate

Life Stage

Species

Nerodia fasciata Crocodylus moreletii Chelydra s. serpentina

Agkistrodon piscivorus

Trachemys scripta Trachemys scripta Akistrodon piscivorous Nerodia fasciata N. taxispilota Nerodia sipedon

Nerodia fasciata

Seminatrix pygaea Nerodia fasciata N. rhombifer N. taxispilota Seminatrix pygaea Nerodia rhombifer Alligator mississippiensis

Table 4.2 Examples of Recent Laboratory and Field Studies with Aquatic (Freshwater) Reptiles

Hopkins et al. 2005 Pepper et al. 2004 Ashpole et al. 2004

Burger et al. 2005 Campbell et al. 2005 Rainwater et al. 2005

Moeller 2004 Tryfonas et al. 2006 Burger et al. 2006

Hopkins et al. 2002

Milnes et al. 2004

Hopkins et al. 2005

Hopkins and Winne 2006

Reference

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Environmental exposure OCs

Metals Mercury Environmental exposure

OCs PCBs OCs Trace elements

Field survey Field survey

Field survey Field survey Field survey

Field survey

Field study Integrated study

Selenium

Field survey

Egg residues Tissue residues

Plasma residues

Tissue residues Tissue residues Blood

Bone composition Tissue residues

Tissue residues

Eggs Adults

Adults

Adults Eggs Adults

Eggs Hatchling Adults Adults Alligator mississippiensis Lepiochelys kempii Caretta caretta Malaclemys terrapin Crocodylits moreletii Agkistrodon piscivorus Nerodia erythrogaster N. rhombifer Trachemys scripta Nerodia sipedon insularium Nerodia s. sipedon Crocodylits moreletii Nerodia fasciata

Alligator mississippiensis

Rainwater et al. 2000 Hopkins et al. 2001

Bishop and Rouse 2000

Burger 2002 Rainwater et al. 2002 Clark et al. 2000

Lind et al. 2004 Keller et al. 2004a, 2004b

Roe et al. 2004

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More often than not, field exposures to environmental chemicals that occur at less than highly acute, but more than no observable effect concentrations, contribute to poor predictions of chemical effects in the natural environment. Mesocosm studies afford an intermediate between field studies and laboratory studies, wherein a few factors considered influential to exposures in field settings are examined. In these intermediate test systems, interactions among environmental factors and chemical contaminants may be studied under controlled situations that are more complex than what can be reproduced in a laboratory. Indeed, exposures more closely approach those observed in open field settings. While field studies more directly relate to the realities of field settings, these studies may not be feasible, and mesocosm studies may provide solutions otherwise dismissed for the sake of more convenient, yet reality-limited laboratory studies. Experiments reliant on mesocosms offer the control of a laboratory study (e.g., allowing examination of a few known factors) but may capture exposures and associated effects under more natural influences (e.g., ambient temperature fluctuations, UV radiation), potentially adding a measure of ecological relevance lacking in laboratory studies. Following a call for an increase in their applications in ecotoxicology (Rowe and Dunson 1994), mesocosm studies have been increasingly used to examine the impacts of chemical contaminants on amphibians (Table 4.1). For example, in their review of the recent ecotoxicological literature focused on amphibians, Boone and James (2005) found 28% of the reviewed studies involved work outside the laboratory, which represents an enormous increase over previous years. Much of this recent work spanned the range of herpetofauna and involved evaluations of chemical effects considered within the context of species life history and endpoints viewed within a multiple stressors framework at various levels of biological organization (e.g., larval amphibian communities experiencing natural stresses of competition for resources, predation, and pond drying). Mesocosms may shed light on how contaminants affect the dynamics of amphibian communities better than laboratory studies, and may better characterize whether contaminants influence amphibians directly or indirectly. For example, in the laboratory environmentally relevant concentrations of carbaryl can negatively impact tadpole behavior (Bridges 1999), yet using mesocosms, adverse effects of carbaryl observed in the laboratory were not necessarily played out in the field. In some studies, carbaryl even appeared beneficial to anurans, since increased body size at metamorphosis was observed (Boone and Semlitch 2002). For salamanders, however, similar exposure conditions were associated with decreased body size at metamorphosis and reduced survival (Boone and James 2003). These contrasting effects appeared to be largely due to carbaryl’s effects on the zooplankton community under the study’s test conditions. Simply stated, carbaryl killed zooplankton. Zooplankton, however, are a food source for salamander larvae, and with the absence or reduced numbers of zooplankton, salamander larvae competed with tadpoles for food resources, for example, algae (Boone and Semlitsch 2002; Mills and Semlitsch 2004). Impacts of carbaryl on tadpoles may also be mediated by indirect effects on the predators rather than directly affecting the tadpoles (Boone and Semlitsch 2003). In laboratory studies, Relyea and Mills (2001) had observed that the presence of a predator was associated with increased carbaryl toxicity. However, in mesocosm studies subsequently focused on the interactions of multiple stressors — pH, predators, and carbaryl — these factors were not associated with outcomes that suggested that interactions would dominate the effects signature displayed in laboratory studies (Relyea 2006).

4.5  Toxicity Assessment: Field Studies Increased awareness of declining amphibian and reptile populations has contributed to their oftentimes being considered “sentinels” of environmental change (e.g., Sparling et al. 2001; Kiesecker et al. 2004). For example, given their dependence on wetlands, some species of amphibians and semiaquatic reptiles may be directly affected by habitat alteration or destruction, by non-pointsource runoff, and by the accumulation of sediments and sediment-bound chemicals that are associated with soil erosion. Physical and chemical alterations of habitat may also be associated directly or indirectly with pathogens (see Carey and Bryant 1995; Crawshaw 2000). Anthropogenic impacts

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on aquatic habitats, and adverse biological effects associated with these habitat alterations have undoubtedly contributed to the retraction of species distributions and decline of herpetofauna populations (Barinaga 1990; Corn 2000). Studies of field populations of amphibians and reptiles may therefore be undertaken to evaluate conditions before the start of some activity that may result in habitat effects (for example, the release of a wastewater discharge, physical disturbance in the riparian zone, flow alterations), or to provide retrospective analyses that indicate the extent of an existing impact, or to evaluate the success of remedial or restoration efforts. For amphibians, a variety of field survey methods have been developed to characterize basic population parameters such as taxa richness and/or diversity (e.g., Heyer et al. 1994; see Table 4.1 for examples). Methods to specifically incorporate herpetofauna into wetland monitoring programs are also available (USEPA 2002). Studies evaluating population level effects of contaminants on reptiles in the field remain limited, with most focused on residue monitoring (Table 4.2). Observational studies and experimental manipulations may also be completed in field settings following various study designs. For example, exclosures or semipermeable enclosures may be used in situ to exclude various influences such as predators or competitors, while allowing most interacting abiotic factors to occur (see Bishop and Martinovic 2000; Linder 2003). While early-stage embryos can be collected from reference sites and placed in areas of concern for evaluating adverse effects in the field, alternative methods may be equally amenable to field studies or integrated fieldlaboratory investigations. While there are drawbacks to field studies (reviewed in Boone and James 2005), many biotic and abiotic influences readily present in the field, but not accounted for in laboratory or mesocosm studies, can be examined. Whole pond manipulations are even less common (but see Boone et al. 2004), as they can be rife with external influences impossible to control or account for statistically. Overall, there appears to be greater focus on monitoring activities in field settings rather than actual experimental manipulations (see Table  4.1). Boone and James (2005) cited only 19 ecotoxicological field studies using amphibians. Most field-only investigations are correlative and require a combination of field and laboratory studies in order to more clearly understand these relationships. Bringing field-collected water into the laboratory for static-renewal exposures lends itself to more controlled experimentation than allowed in the field, but can be cumbersome when large volumes of water are required to satisfy test conditions. Another option is the use of semipermeable membrane devices (SPMDs) to sequester certain contaminant types from an environment, which are then evaluated under more controlled laboratory settings (Bridges and Little 2003). These studies have been a useful first step in situations where it is desirable to determine whether contaminants are impacting amphibian populations without having to undergo costly chemical analyses on water samples. For example, Davidson et al. (2001, 2002) observed that a number of amphibian species with declining populations occurred upwind from agricultural lands in the central valley in California. Using SPMDs, Bridges and Little (2005) were able to determine the presence of chemical contaminants within these areas, and that reduced growth and development of native tadpoles occurred when SPMDs were evaluated in laboratory toxicity tests. Sparling et al. (2001) had noted inhibition of cholinesterase in field-caught frogs collected from ponds similarly situated to those observed by Davidson et al. (2001, 2002). A range of integrated field and laboratory studies have focused on various environmental chemicals and their potential risks for herpetofauna in various managed landscapes, as illustrated by Hayes et al. (2003). In this study, leopard frogs were sampled at various locations across the United States, and the incidence of intersex individuals was correlated with concentrations of the herbicide atrazine measured at these sampling locations. These findings from field-collected samples were further addressed in controlled laboratory investigations that indicated that low-concentration exposures to atrazine can generate intersex individuals (Hayes et al. 2002). Similarly, Fort et al. (1999) had earlier applied similar approaches for investigating the role of environmental chemicals in explaining observations of hind limb deformities in anurans collected from ponds in Minnesota and Vermont. In these studies, Fort et al. (1999) used a combination of laboratory FETAX assays of water and sediments to characterize causal relationships between chemical exposure and the

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high incidence of amphibian malformations. While the role of integrated field and laboratory studies may involve any organism amenable to such studies, the herpetofauna (especially, amphibians) provide witness that the integration of field and laboratory studies combines the strengths of both approaches in assessing risk of chemical stressors and reducing uncertainties potentially influencing management actions.

4.6 Biomarkers of Exposure and Effect in Amphibians and Reptiles Biomarkers have been defined as biochemical, physiological, and histological endpoints that can be used to evaluate exposure to or effects of chemical stressors (Huggett et al. 1992). Some authors apply a more general definition that includes morphological alterations, genetic effects, behavioral parameters, and tissue residue levels (Walker et al. 2001). Venturino et al. (2003) provided an extensive review of biomarkers that have been used to indicate contaminant effects in anuran amphibians, and additional examples for both amphibians and reptiles are presented in Table 4.3. An expected advantage of evaluating biochemical and physiological parameters in organisms exposed to chemical contaminants is that these endpoints may be influenced by a stressor before responses are seen at the whole organism or population level. As such, biomarkers may provide an early indication of contaminant exposure and/or effects (Newman and Unger 2003). It may also be possible to compare values of parameters from a contaminant-exposed population with reference values to indicate the influence of chemical exposure on the general health status of the organisms. While a lack of suitable reference data may limit this application for herpetofauna (Henry 2000), this issue exists for other wildlife species and may be addressed by comparison to a suitable reference population. When interpreting the results of such evaluations, it is important to consider the influence of potentially confounding factors such as sex, developmental stage, and season (Rie et al. 2000; Venturino et al. 2003). In addition, the influence of temperature on physiological processes in poikilotherms like amphibians and reptiles is an important consideration when comparing biomarkers between different populations. For example, Johnson et al. (2005) found a significant difference in acetylcholinesterase activity (a common biomarker of exposure to carbamate and organophosphorous pesticides) between 2 groups of Pacific tree frogs (Hyla regilla) that had been raised from tadpoles under different temperature regimes. Basic differences in biochemical and physiological characteristics between animal groups can also affect the utility of some variables for indicating contaminant exposure. For example, induction of liver enzymes that are part of the mixed-function oxidase (MFO) system has been used to indicate exposure to a range of organic chemicals (including pesticides) in a number of vertebrates, but reduced activity of these enzymes in amphibians may limit the use of this parameter as a biomarker for this group (DeGarady and Halbrook 2003; Venturino et al. 2003). Biomarkers have been evaluated in herpetofauna in both laboratory and field settings (e.g., Overmann and Krajicek 1995; Vogiatzis and Loumbourdis 1999; Rie et al. 2001; Keller et al. 2004a, 2004b), and may be useful for bridging studies that compare the exposure scenarios. However, a potential weakness regarding the use of these variables for risk assessment is the often poor understanding of the ecological relevance of observed responses in biomarkers. For example, Widder and Bidwell (2006, 2008) observed up to a 43% reduction in cholinesterase activity in southern leopard frogs (Rana sphenocephala) exposed to the organophosphate pesticide chlorpyrifos, but did not observe any effect on survival, growth, or swim speed. Studies such as those by Sparling et al. (2001) that evaluated possible links between cholinesterase enzyme activity, pesticide residue levels, and population status of anuran amphibians in the Sierra Nevada Mountains of California, and DeGarady and Halbrook (2003), which evaluated a biomarker of organic contamination along with abundance and richness of amphibian populations from sites subject to long-term PCB contamination, are important for indicating physiological and biochemical variables that may best be linked to responses at higher levels of organization. However, there remains a need for studies that more clearly demonstrate how a response to chemical exposure that originates at the subcellular level will

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Genetic markers

Immunologic markers

Other liver enzymes and carbohydrate stores Endocrine

Numerous, including nonspecific cytotoxic cells (NCCs) and macrophages DNA strain breakage sister chromatid exchange

Various chemicals

Low pH

Various chemicals

Effects on sex hormones — gonad morphology, reproductive hormone levels, vitellogenin levels

Effects on thyroid hormones — thyroxine (T4) and triiodothyronine (T3) levels

Copper

Polynuclear aromatic hydrocarbons, other organochlorine compounds

Lead

Lead Organophosphate pesticides

Initiating Chemical Stressor(S)

Aminotranferases and glycogen

Intermediate metabolites or degradation products of heme synthesis, e.g., porphyrins and aminolevulinic acid dehydratase (ALAD) Mixed-function oxidases and associated enzymes

Indicators of synthesis disruption

Xenobiotic metabolism

Inclusion bodies in kidney Esterases, reductases, e.g., acetylcholinesterase

Specific Marker

Histological effects Enzyme inhibition

General Biomarker Category

Rana catesbiana Bufo boreas

R. nigromaculata Chinemys reevesii Rana esculenta Alligator mississippiensis Rana pipiens

Xenopus laevis Rana pipiens

Rana ribunda

Bufo arenarum Rana ridibunda

Rana ridibunda Hyla regilla Rana sphenocephala Rana spp. Nerodia rhombifer Agkistrodon piscivorus Trachemys scripta Bufo arenarum

Species

Table 4.3 Examples of Recent Biomarker Studies Conducted with Amphibians or Aquatic (Freshwater) Reptiles

Wirz et al. 2005 Tverdy et al. 2005

Yang et al. 2005 Tada et al. 2004 Mosconi et al. 2005 Gunderson et al. 2002 Vatnick et al. 2006

Hayes et al. 2002 Hayes et al. 2003

Venturino et al. 2001 Kostaropoulos et al. 2005 Gunderson et al. 2004 Papadimitriou and Loumbourdis 2005

Arrieta et al. 2004

Loumbourdis 2003 Sparling et al. 2001 Widder and Bidwell 2006, 2008 Van den Brink et al. 2003 Clark et al. 2000

Reference

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ultimately extend to the population and community, and also how environmental variables such as temperature influence these relationships. Several studies have examined the influence of contaminants on selected suborganismal parameters in reptiles, with most focus on aquatic species (e.g., Lamb et al. 1995; Overmann and Krajicek 1995; Ulsh et al. 2000; Willingham and Crews 2000; Sanchez-Hernandez 2003; Keller et al. 2004a, 2004b; Tada et al. 2004). In their review of the ecotoxicology of metals in amphibians and reptiles, Linder and Grillitsch (2000) included a discussion of the available literature that examined biochemical effects on reptiles, and Portelli and Bishop (2000) provided a similar overview for reptiles (mostly turtles) exposed to organic contaminants. Meyers-Schöne and Walton (1994) also discussed biochemical and histopathological responses to stress in their review on the use of turtles for monitoring chemical contaminants. Due to their generally larger body size, certain aquatic to semiaquatic reptiles may be more amenable to studies of biomarkers than amphibians, since reptiles generally offer greater tissue mass and/or blood volume for analyses. There may also be a greater possibility of extracting blood or tissue samples from reptiles without having to kill the organisms, an important consideration when working with low-density populations or threatened or endangered species. If tissue residues accumulated through bioconcentration and bioaccumulation are considered biomarkers of exposure, studies with reptiles are probably more common than those with amphibians (Table 4.2). Furthermore, aquatic species are more frequently reported in the literature when our focus resolves on tissue residue studies. Sparling (2000) stated that, when considering the taxonomic diversity of reptiles, a disproportionate amount of research had been conducted on turtles and tortoises, with much of this work focused on metals, chlorinated pesticides, and polychlorinated biphenyls (PCBs). Since that publication, there has been an increase in the number of papers that studied the effects of these contaminants (Chapter 1, this book). Reviews of tissue residue studies for both amphibians and reptiles can be found in Sparling et al. (2000 and various chapters of this book), and Meyers-Schöne and Walton (1994) have discussed residue studies with turtles. Campbell (2003) assembled, reviewed, and summarized available organic, inorganic, and radionuclide contaminant accumulation and effects studies for crocodilians and characterized data gaps in order to promote their future inclusion in environmental contamination studies and ecological risk assessments. Reptile eggs can accumulate contaminants from both maternal transfer and soil during the incubation period and may serve as indicators of chemicals that are bioavailable. For example, Pepper et al. (2004) examined the exposure of crocodiles to organochlorine (OC) pesticides using the chorioallantoic membrane (CAM) of reptile eggs. CAM is critical to embryonic development when it functions in gas exchange, nutrient transport, and waste storage for the developing embryo. As a nonlethal biomarker, CAM serves as a noninvasive indicator of exposure, since it remains with the eggshell after hatching, and has been successfully used to examine contaminant exposure and predict chemical concentrations in multiple species of birds and egg-laying reptiles (see, e.g., Pastor et al. 1996; Cobb et al. 1997). In their study, Pepper et al. (2004) found OC burdens in crocodile CAMs confirmed contamination of eggs, which suggested that exposure to females and embryos in ovo had occurred. Recent work by Unrine and Jagoe (2004), Fagotti et al. (2005), Keller et al. (2005), Gardner et al. (2006), and Tryfonas et al. (2006) further illustrates the continued interest in studying contaminant residue levels in tissues of amphibians and reptiles in both controlled laboratory experiments and field monitoring studies. As reviewed by Hayes (2000) and Guillette (2000), amphibians and reptiles have played an important role as indicators of endocrine-disrupting chemicals in aquatic systems. Most work to date has focused on effects related to sex determination, with biomarkers including gonadal morphology and/or circulating levels of plasma hormones or specific proteins (e.g., Noriega and Hayes 2000; Shelby and Mendonca 2001; Hayes et al. 2003). Many egg-laying reptiles may be particularly suited for studies of chemicals that affect sex hormones, since they display temperature-dependent sex determination that may facilitate manipulation of sex ratios in experimental animals to more clearly evaluate chemical effects (Crews et al. 1995; Newman and Unger 2003). Recent studies have

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also indicated a significant role for herpetofauna in the evaluation of chemicals that disrupt the thyroid axis, since thyroid hormones are important for initiating metamorphosis in amphibians or egg hatching in reptiles (Brasfield et al. 2004; Furlow and Neff 2006; Tata 2006). The study of physiological energetics is another tool that deserves mention under the general heading of biomarkers. A clear advantage of studying variables related to energy balance is their direct ecological links to individual, population, and community levels of biological organization (Congdon et al. 2001). Rowe et al. (2003) discuss energetics as it relates to larval, juvenile, and adult stages of anuran amphibians and the role chemical stressors may play in increasing maintenance costs and decreasing energy available for growth. Amphibians may be particularly sensitive to factors that deplete energy reserves during metamorphosis, since feeding may cease during this time due to reorganization of the mouthparts and digestive system (Rowe et al. 2003). Reptilian eggs may also serve as valuable models to study the energetic effects of chemical stressors, since development of the embryo relies entirely on internal yolk stores. As such, this dependency likely increases exposure in ovo, since contaminants may pass across the eggshell in association with imbibed water (Moeller 2004).

4.7 Recurring and Emerging Issues: Future Challenges for Toxicologists Studying Amphibians and Reptiles 4.7.1 Recurring Issues in the Ecotoxicology of Amphibians and Reptiles Surrogate species are widely used in the field of ecotoxicology, particularly when the species of primary concern is threatened, endangered, or simply hard to come by. The herpetofauna are rife with members that are characterized by sparse to nonexistent ecotoxicological information, and often species are relatively poorly characterized with respect to their life histories (e.g., caecelians, amphisbaenids). As demonstrated in Chapter 1 of this book, most ecotoxicological studies have focused on 4 genera of amphibians, including Rana, Bufo, Ambystoma, and Xenopus. These data gaps present problems regarding species sensitivity to environmental chemicals and assessing threats that contaminants may pose. Because many species of amphibians and reptiles continue to display declines in their populations, the use of surrogate species has a particularly high demand within this group. Yet, many factors must be considered when using surrogates, as the herpetofauna present outward appearances that belie the diversity of amphibians and reptiles in terms of life history strategies. For example, most ecotoxicological studies focused on amphibians rely on the aquatic stage of the biphasic life cycle that is typical of many North American species. But, amphibians include members that are strictly fossorial (e.g., caecelians), some are solely aquatic (e.g., sirens, hellbenders), and others bypass the process of metamorphosis in favor of direct development (e.g., some plethodontid salamanders). Amphibian eggs can develop in water (e.g., ranging from waters of lotic or lentic habitats to pools lying within the folds of a bromeliad leaf), under logs, and even within the vocal sacs or along the backs of parent frogs. Consequently, data collected from species exhibiting typical biphasic life cycles may not be as relevant for these alternative life history strategies, especially when long-term exposure to less than acutely lethal concentrations is considered. Similarly, within those species presenting a biphasic life cycle, a range of life history strategies has been observed. Aquatic amphibian species differ from one another in the length of their larval period; for example, some desert-dwelling species of toad complete metamorphosis in 8 days, while bullfrogs in North American may take 3 years to go undergo the process. Similarly, neotenic salamanders display facultative metamorphosis, and depending on environmental conditions, these species oftentimes remain aquatic throughout their lives. The longer the larval period, exposures to waterborne contaminants will be increased, and even relatively resistant species may express adverse effects associated with prolonged dependencies on aquatic habitats. Other life history attributes similarly affect exposure and remain a recurring source of uncertainty when reliance on

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surrogate species data is incorporated into the risk assessment process for amphibians and reptiles. For example, dietary exposures differ among herpetofauna, as exemplified by anuran larvae generally being filter feeders, while most urodeles are carnivorous. These differences in early life stages undoubtedly affect exposure and consequent effects. Reptiles present similar challenges with respect to life history strategies, especially regarding their poorly developed reproductive toxicity data for any species. Reproductive strategies are widely divergent among the reptiles, which contributes to the problem of developing suitable surrogates for assessment purposes. Reptiles can be oviparous, viviparous, or ovoviviparous, which necessarily complicates development of test systems focused on critical life stages involving in ovo and in utero exposures, as well as maternal transfer of chemicals; for example, in viviparous species, maternal transfer may be the only route of exposure during development, but in oviparous reptiles maternal transfer may be supplemented by uptake of environmental chemicals from the surrounding soil or matrix of nest materials. As noted by Linder et al. (Chapter 5, this book) in their overview of the physiological ecology of herpetofauna, amphibians and reptiles differ markedly in their adaptations to the environment, differences that inevitably influence exposure in these animals. For example, in contrast to amphibians, reptiles have keratinous skin that tends to limit dermal uptake of water and water-soluble environmental contaminants. Life history attributes linked to reproduction also influence exposure, as exemplified by differences in potential for maternal transfer of chemicals between mother and offspring. Reptiles also tend to produce fewer offspring per reproductive effort and are longer-lived than amphibians, which are life history attributes that complicate development of standardized toxicity tests. Indeed, the diversity of the herpetofauna is no less, and likely exceeds, that of the charismatic megafauna that frequently dominate the ecological risk assessment process. As with any surrogate species, if species of amphibians and reptiles amenable to routine toxicity assessments become available in the future, caution must be applied to minimize uncertainties associated with broad generalizations developed from these select few whose life history attributes ensure their being amenable to laboratory manipulation but inevitably may mean they are far from representative. Indeed, the good of the few, or the one, may not outweigh the good of the many.

4.7.2 Emerging Chemical Contaminants Perhaps foreshadowed by observations of endocrine-disrupting chemicals in the environment (see Hayes 2000; Guillette 2000), the past 5 to 10 years have been characterized by the growing recognition that a number of chemicals not previously considered as contaminants are present in the environment on a global scale (USGS 2006), especially in surface waters, sediments, and water treatment residuals (e.g., sewer sludge and biosolids). Future investigations will undoubtedly lead to expanded discovery of these “emerging chemical contaminants” in terrestrial and wetland habitats. These emerging contaminants include hormones, pharmaceuticals, personal care products, and other organic compounds that are frequently derived from municipal, industrial, and agricultural sources. See Kolpin et al. (2002), Barnes et al. (2002), and Chapter 15, this book, for comprehensive lists. Yet undiscovered are environmental derivatives associated with the developing nanotechnology industry and intentional, coincidental, or accidental release of these materials to the environment. Regardless of their form, the discovery of these contaminants, largely facilitated by the development of analytical techniques that allow their detection, has met with increased concern for their effects on humans and ecological receptors in the environment. Sanderson et al. (2004) characterized 4 broad classes of pharmaceuticals found in freshwater (antibiotics, antineoplastics, cardiovascular drugs, and reproductive hormones) and used quantitative structureactivity relationships to predict that nearly one-third of all drugs could be very toxic to aquatic organisms. Not surprisingly, however, empirical toxicological data for these chemicals are largely absent, although recent studies anticipate future efforts, particularly for the herpetofauna (primarily, amphibians). Smith and Burgett (2005) reported no effects on growth and variable effects on other measures of biological activity levels in Bufo americanus tadpoles exposed to individual

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treatments of 3 common organic wastewater contaminants (acetominaphen, an antipyretic; triclosan, an antimicrobial agent; and caffeine, a stimulant). Studies with fish have indicated that wastewater contaminants with low estrogenic activity could have a combined potency that leads to observable biological effects, and that longer-term chronic tests are most appropriate to elucidate these effects (Rodgers-Gray et al. 2000; Thorpe et al. 2003; Sumpter and Johnson 2005). Needless to say, where amphibians and semiaquatic reptiles fit into the range of responses must be further characterized with test systems that provide exposure to realistic test concentrations and exposure durations. Beyond the pharmaceuticals, other chemicals have recently emerged as contaminants of ecological concern, in part because of improved analytical techniques that have allowed for their detection at environmental concentrations. For example, perfluorooctanesulfonate (PFOS) and related perfluorinated compounds were historically used in numerous industrial and consumer products because of their capacity to repel both water and oil. While no longer manufactured in the United States, PFOS and associated compounds are increasingly being considered chemicals of concern because of their persistence and widespread distribution. In toxicity studies with northern leopard frogs, Ankley et al. (2004) observed bioaccumulation of waterborne PFOS and effects on growth and time to metamorphosis. However, based on the few studies completed, these workers posited that anurans do not appear to be exceptionally sensitive to PFOS in terms of either direct toxicity or bioconcentration potential. Before their early dismissal, however, the contaminant’s comparative toxicity and its bioconcentration and bioaccumulation potential among aquatic test species must be more adequately addressed within the context of risks linked to multiple stressor exposures commonly encountered in field settings.

4.7.3 Ecotoxicology and Multiple Stressors As Burkhart et al. (2003, p 111) observed, “Contaminants typically occur in aquatic and terrestrial environments as complex mixtures of natural and anthropogenic origin, yet the evaluation of the effects of chemical contaminants on amphibians is still primarily based on exposure to single compounds under highly controlled conditions.” While chemical-by-chemical evaluation supports a relatively uncluttered understanding of individual compounds within the context of species, life stage, dose, and mode of action, it falters within an ecological context, stemming from our inability to adequately address chemical interactions, either among chemicals of the mixture or between chemicals of the mixture and other components of the environmental matrix in which it occurs. In contrast to the laboratory environment, field settings are highly dynamic and reflect exposures to numerous chemical and nonchemical stressors. There are many scenarios for potential acute, chronic, and pulsed exposures in amphibians across various life stages and in reptiles challenged by environmental chemicals across their range of habitats. Data are beginning to accumulate suggesting that the detrimental effects of aquatic contaminants on amphibians and reptiles are underestimated using the approaches commonly applied in ecotoxicology investigations (Fort et al. 1999; Mann and Bidwell 1999; Sparling et al. 2000a; Relyea 2004, 2005). For example, Johnson and Sutherland (2003) discussed the importance of considering multiple environmental stressors, including contaminants, when investigating the interrelationships between the causative agents of trematode infections that might induce limb deformities in amphibians and co-occurring chemical stressors. Clearly, multiple stressor considerations are not unique to herpetofauna, but could influence chemical response in any ecological (or human) receptor. However, the biphasic nature of the common amphibian life cycle and the occurrence of semiaquatic reptiles at the land-water interface inevitably suggest that nonchemical stressors may be particularly important components of exposure-response scenarios, especially when spatial heterogeneity of habitat yields a gradient of exposure conditions. Physical habitat alteration such as sedimentation associated with soil erosion in areas of disturbance, and interactions between those physical stressors and environmental chemicals have frequently been identified as major factors in multiple stressor exposures (e.g.,

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Linder et al. 2002). Also, temporal components have frequently been identified as major factors influencing exposure. For example, seasonal changes in wetlands may be management-critical factors in reducing or mitigating risk associated with environmental chemicals. Spring runoff and autumn dry-down will affect environmental concentrations of agrichemicals in wetlands and seasonal ponds. At least 2 synthesis publications focusing on multiple stressor effects in amphibians (Linder et al. 2003a, 2003b) highlight the importance of these issues.

4.8 Summary From the earlier overview of the ecotoxicology of amphibians and reptiles by Sparling et al. (2000), it became apparent that the available literature through 1998 was relatively sparse compared to publications focusing on wild mammals, birds, and fishes. The chemical contaminant literature available for amphibians had been haphazardly developed over the preceding 25 to 30 years and had focused on metals and chlorinated organic chemicals to yield data sufficient for speculative analyses of risks for these environmental chemicals. In contrast, much of the literature for reptiles focused on compiling tissue residue values for free-ranging animals. The past 10 to 15 years have yielded an increase in research on herpetofauna and the effects of environmental chemicals on these taxa. An updated search focused on herpetofauna as key elements in aquatic toxicology studies yielded over 700 publications since 1998 (see Chapter 1, this book), a number far outpacing annualized counts reported in that earlier publication (Sparling et al. 2000). Today, amphibians are more commonly considered in toxicity assessments, whether implemented as detailed for FETAX in ASTM E-1439 or as any of various alternatives. Still, much work remains to be initiated and completed, if the herpetofauna are going to be sufficiently represented in the discipline of ecotoxicology. There is little argument that amphibians and reptiles are sensitive to environmental chemicals (Birge and Black 1977; Cowman and Mazanti 2000; Sparling 2000; Hayes 2000; Ouellet 2000). Fortunately, the available cross-species data from single-chemical toxicity tests do not suggest that the herpetofauna are a priori more sensitive or more resistant than any other species (McCrary and Heagler 1997), but the data available for such comparisons are not sufficient to conclude that existing “safe levels” for single chemicals are protective across the range of species and habitats in which the herpetofauna occur (see Vaal et al. 1997). At present, uncertainties relative to risks associated with exposures in aquatic, wetland, and terrestrial systems vary from adequate, for example, for simple evaluations of the toxicity of some metals, to absent for the vast majority of chemicals that herpetofauna encounter in the field. Work with reptiles still lags behind that of amphibians, especially with respect to questions focused on habitats, interactions, and life history. Their roles in exposures of reptiles to environmental chemicals remain largely undiscovered. In addition, environmental contamination studies are relatively limited and are not available for many reptiles. For example, more than half of the 23 crocodilians remain unstudied, and when these reptiles have been considered in field studies, efforts focused on accumulation and effects of mercury and endocrine-disrupting chemicals (EDCs) on American alligator in Florida (see, e.g., Guillette et al. 1994, 1995; Heaton-Jones et al. 1997). Campbell (2003) indicated that the effects of EDCs on crocodilians are not confined regionally and probably occur in many parts of the world, especially in developing tropical areas where organochlorine pesticides are more widely used. Similarly, effects associated with inorganic contaminants such as mercury and other metals are poorly characterized in these species. At present, any aquatic, sediment, or soil benchmark values are not supported by the scant empirical data that are available for these organisms. Although long overlooked and consistently undervalued, amphibians and reptiles have recently gained increased recognition as critical components within many ecosystems. That heightened awareness among resource managers and members of the research community, however, must be matched by increased efforts to address data gaps in our existing knowledge of chemical toxicity to the wide range of species in these vertebrate groups. More importantly, the interrelationships of these animals with other ecosystem attributes and other physical and biological stressors must be characterized to

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enable amphibians and reptiles to better serve as indicators of habitat quality and ecosystems at risk. Indeed, for aquatic habitats such as wetlands of various types, indigenous amphibians may be more important to evaluating system sustainability than presently appreciated. Similar observations apply to reptiles where their unique roles and habitat dependencies (e.g., deserts throughout their various categories) often outpace those of wild mammals and birds. The herpetofauna must receive increased attention in future environmental research, particularly as that relates to enhancing our understanding of their ecotoxicology and the role that long-term, low-level chemical exposures play in their future. Failure to do so may ultimately serve as a harbinger of our shared loss.

Acknowledgments The authors extend a hearty thank you to the biologists whose efforts in the laboratory and field helped bring amphibians and reptiles to the attention of regulators and natural resource managers long focused on charismatic megafauna. Although our work has only begun, everyone’s efforts are greatly appreciated.

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Physiological Ecology of Amphibians and Reptiles Natural History and Life History Attributes Framing Chemical Exposure in the Field Greg Linder, Brent D. Palmer, Edward E. Little, Christopher L. Rowe, and Paula F.P. Henry

If we are what we eat, then we are fast, cheap, and easy. — Anonomyous

Contents 5.1 Physiological Ecology and Exposure to Chemicals in the Environment.............................. 107 5.1.1 Brief Overview of Physiological Energetics.............................................................. 108 5.1.2 Physiological Ecology, Exposure Models, and Food Chains.................................... 110 5.2 Pathways of Contaminant Exposure for Amphibians and Reptiles....................................... 111 5.3 Physiological Ecology of Amphibians and Reptiles: Natural History and Life History Attributes Influencing Exposure............................................................................................ 111 5.3.1 Dietary Exposures, Gastrointestinal, and Digestive Physiology............................... 112 5.3.2 Thermoregulatory, Osmoregulatory, and Excretory Physiology............................... 115 5.3.3 Dermal and Percutaneous Exposure, Respiratory Physiology, and Gas Exchange.................................................................................................................... 119 5.4 Endpoints Commonly Linked to Chemical Exposures to Amphibians and Reptiles in Laboratory and Field......................................................................................................... 122 5.4.1 Growth....................................................................................................................... 122 5.4.2 Reproduction and Endocrinology.............................................................................. 124 5.4.2.1 Reproduction and the Environment............................................................ 125 5.4.2.2 Pineal Gland............................................................................................... 125 5.4.2.3 Hypothalamus and Pituitary....................................................................... 125 5.4.3 Female Reproduction................................................................................................. 126 5.4.3.1 Vitellogenin................................................................................................. 126 5.4.3.2 Ovarian Structure and Function................................................................. 126 5.4.3.3 Reproductive Strategies.............................................................................. 127 5.4.4 Male Reproduction.................................................................................................... 128 5.4.4.1 Testis Structure and Function..................................................................... 128 5.4.4.2 Fertilization and Copulatory Organs.......................................................... 128

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5.5 Reproductive Ecology............................................................................................................ 129 5.5.1 Parental Care............................................................................................................. 129 5.5.2 Offspring Survival..................................................................................................... 129 5.5.3 Lifespan and Exposure.............................................................................................. 130 5.6 Development.......................................................................................................................... 130 5.6.1 Sex Determination..................................................................................................... 130 5.6.2 Metamorphosis.......................................................................................................... 132 5.6.3 Endocrine-Disrupting Compounds............................................................................ 134 5.7 Behavior................................................................................................................................. 135 5.7.1 Sensory Organs.......................................................................................................... 135 5.7.2 Locomotion and Foraging......................................................................................... 136 5.7.3 Amphibians and Reptiles: Entanglements of Chemical Exposures, Foraging, and Feeding Habits.................................................................................................... 136 5.8 Biomarkers, Metabolism, and Development of Energetics-Based Tools............................... 139 5.8.1 Specific Dynamic Action........................................................................................... 141 5.9 Interactions of Chemicals with Physiological and Environmental Factors........................... 142 5.9.1 Ultraviolet Radiation................................................................................................. 143 5.9.2 Stress.......................................................................................................................... 145 5.9.2.1 Hibernation................................................................................................. 146 5.9.2.2 Freeze Tolerance......................................................................................... 147 5.9.2.3 Estivation.................................................................................................... 147 5.10 Physiological Ecology and Multiple Stressors: Developing a Common Currency to Evaluate Chemical Exposures to Amphibians and Reptiles in Field Settings...................... 147 5.10.1 Research Needs: The Next 10 Years and Beyond...................................................... 148 Dedication....................................................................................................................................... 149 References....................................................................................................................................... 149 With publication of the first edition of Ecotoxicology of Amphibians and Reptiles (Sparling et al. 2000a), the environmental toxicology community was introduced to historically undervalued animal classes whose contaminant-related literature was highly diffuse and relatively sparse. In contrast to mammals and birds, throughout the short history of ecotoxicology, “one-stop” information resources for herpetofauna were few in number and frequently difficult to obtain. More frequently, these data and information resources were nonexistent. Sparling et al. (2000a, 2000b), then, achieved the goals of making the science and resource management communities aware of the role that amphibians and reptiles should play in ecotoxicology, and identified research necessary to address the shortfall in ecotoxicological information critical to resource managers working in conservation programs focused on these animals. Now early in the 21st century, the science of ecotoxicology has entered its adolescence, and an update to that first edition is amply warranted. However, for much of the original edition’s coverage, a simple update would entail additional materials that would preclude a thrifty second edition. Hence, the present chapter does not simply replace the works of Henry (2000) and Palmer (2000) on the physiology of amphibians and reptiles, respectively, but focuses on physiological traits of amphibians and reptiles that place constraints on fitness when exposed to contaminants in the field. In short, our primary focus resolves on the physiological ecology of amphibians and reptiles, particularly with reference to the interactions between animals and their environment as they affect exposure. Whereas our primary focus resolves on the nexus of physiological ecology and ecotoxicology of the herpetofauna, it would have been presumptuous of us had we thought that characterizing a fully developed state of the science was easily attainable in our single chapter, particularly given the scant literature available for the herpetofauna. Rather, we present a few examples of how studies regarding the ecotoxicology of the herpetofauna could benefit from the mind-set of a physiological

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ecologist when characterizing exposure, and we look toward traditional physiological literature to shape the context of our interpretations. Indeed, Henry (2000), Palmer (2000), Feder and Burggren (1992), Duellman and Trueb (1994), and the classic references of Noble (1931) and Gans (Gans and Pough 1982a, 1982b; see Dudley et al. 2006) should not be off-loaded to secondhand book stores or lost to the obscure bowels of reference libraries. And, references such as McNabb (2002), Prosser (1991a, 1991b and earlier editions), Schmidt-Nielson (1997 and earlier editions), and Kasarov and Martínez del Rio (2007) should become common companions to these contemporary classics in the libraries of current and future generations of ecotoxicologists. Indeed, these references remain vital source materials for research ecotoxicologists and the wide spectrum of biologists studying amphibians and reptiles.

5.1 Physiological Ecology and Exposure to Chemicals in the Environment Simply characterized, physiological ecology concerns the biophysical, biochemical, and physiological traits that have evolved in response to chemical and physical factors in the environment. In concert with an individual’s behavioral and morphological adaptations, these processes mediate spatiotemporal interactions with other organisms in shared habitat. Frequently, an energetics-based approach is applied to the study of physiological ecology, which potentially affords the ecotoxicologist a wide variety of tools for analysis of exposure and effects linked to chemical stressors. Energetics concerns the flow of energy and material within a system at any level of biological organization — molecular, cellular, organismal, population, community, or ecosystem. Applying an energetics-based framework for evaluating exposure and effects enables the use of a “common currency” readily applicable to evaluations of multiple stressors and capable of easing the analysis of chemical stressor effects within and among levels of organization. This framework potentially reduces uncertainty associated with extrapolations traditionally made from individual to population levels of organization, and further to landscapes and ecosystems. As suggested by Linder, Lehman, and Bidwell (Chapter 4, this volume), physiological ecology, and more specifically physiological energetics, deserves greater attention in the evaluation of exposure and effects of chemical stressors in amphibians and reptiles. Congdon et al. (2001) and Rowe et al. (2003) clearly identified that a physiological energetics framework provides a foundation for evaluating chemical stressors, since the focus on variables related to energy flow directly links ecological attributes of individuals, populations, and communities. Depending on the level of biological organization and the spatiotemporal setting for the analysis, opting for a physiological ecologist’s view of energy and material flows provides for rigorous evaluation of the effects of multiple stressors on amphibians and reptiles. For example, Rowe et al. (2003) discussed physiological energetics as it relates to larval, juvenile, and adult stages of amphibians and the influences of chemical stressors on maintenance costs and, subsequently, growth. Similarly, in reptiles early developmental stages may serve as valuable models to study energetic effects associated with exposures to chemical stressors (e.g., embryos may be exposed to chemicals due to maternal transfer during vitellogenesis and from uptake of water from contaminated soils in the nest; Moeller 2004). Nutritional, behavioral, and energetic interactions influence exposure to environmental chemicals and may dramatically affect risks to wildlife. Behavioral interactions influencing exposure were observed early in ecotoxicological investigations (e.g., Steele et al. 1989; Strickler-Shaw and Taylor 1990, 1991; Taylor et al. 1990; Steele et al. 1991 on amphibians), but nutritional and energetic interactions that modify exposure remain relatively poorly characterized in most ecotoxicological contexts, as evidenced by these topics being absent or practically so in references such as Hoffman et al. (2003), Newman and Unger (2003), Newman and Clements (2007), Schüürmann and Markert (1997), and Walker (2001). For discussions focused on amphibians and reptiles, the technical aspects of these interactions are even more scant than for other vertebrates. Indeed, few

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studies have characterized their roles in influencing exposure and mediating biological effects that may dramatically affect the risk to herpetofauna. The ecological consequences of these biological processes may be associated with responses that are as significant as toxicity linked directly to chemical exposure. Our understanding of exposures linked to whole animal interactions with environmental substrates (e.g., dermal uptake of chemicals from soils) is also limited, although an increasing number of studies are addressing chronic exposures in sediments, soils, and other matrices (James 2003; James et al. 2004a, 2004b). Routine consideration of nutritional, behavioral, and energetic processes would provide a biological and ecological context to ecological risk assessment; however, these factors have often been neglected in the analysis and interpretation of chemical risks to wildlife, especially amphibians and reptiles.

5.1.1 Brief Overview of Physiological Energetics The flow of energy and material in an animal is illustrated in Figure 5.1. The dominant practice today for evaluating hazards and risks focuses on materials, or more specifically chemicals within and transfers between a system’s compartments. Food consumption is routinely measured or estimated in the traditional exposure model for terrestrial and semiaquatic vertebrates. Yet, the evaluation process could easily be extended in research studies undertaken beyond a screening level application. Combined with measures of tissue residues and material inputs, characterization of intake energy (IE) would provide the basis for more comprehensive studies of energy allocation and material partitioning among “loss compartments” (urinary and fecal production) and “storage compartments” (tissues). At present, however, little thought is given to energy-materials interrelationships or differential expenditure of energy when challenged by multiple factors, be they chemical, physical, or biological. While the process summarized in Figure 5.1 and detailed elsewhere tracks energy and materials (for example, see Karasov and Martínez del Rio 2007), the compartments characteristic of the biological process are conveniently identified by their various forms of “energy,” since the physiological ecologist’s focus often resides in organismal “currency” (e.g., kilocalories) and its relationship to materials. If we follow conventional leads and focus on food chain as our tool to evaluate exposure in wildlife, the role of energy flow in material transfers becomes an integral component of the analysis. Here, IE (generally measured in calories or joules determined by complete oxidation of

Fecal Energy Intake Energy

Digestible Energy Stored or Tissue Energy

Extended Metabolic Reproduction Activity (e.g., foraging, competitive interactions)

Maintenance

Urinary Energy Gaseous Energy Metabolizable Energy Heat Energy Net Energy

Basal Metabolism

Figure 5.1  The conceptual model of energy and material flow provides a physiological energetics framework for evaluating exposure with traditional food chain models.

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material in a bomb calorimeter) represents the total energy in materials consumed (e.g., forage items plus substrates such as soil and sediment). Most often, this would be estimated as a function of gross energy (energy per unit mass) and foraging rate (mass consumed per unit time). Not surprisingly, gross energy and nutritional content of feed materials vary widely from season to season and from year to year; hence, IE will likely vary as well. Additionally, thermodynamic theory dictates that individuals cannot be 100% efficient in material and energy transfer and utilization. Therefore, the amount of food and energy not used by the animal must be determined, if a realistic accounting of food constituents — be they nonnutritive constituents, nutrients, or chemical toxicants — is desired. Fecal energy (FE) characterizes the material and energy content of feces that can be measured in samples collected in field or controlled laboratory feeding studies. As such, digestible energy (DE) is relatively straightforward in its derivation, being the difference between ingested energy and fecal energy (IE – FE). This estimate is more accurately referred to as “apparent digestible energy,” since contributions to fecal energy are derived from multiple sources, including the microbial flora of the gut, intestinal secretions from the gastrointestinal tract, and sloughed intestinal lining. Determination of true DE will likely remain a research question of nutritional ecologists, yet opting for a physiological ecologist’s mind-set seems a preferred basis for research efforts targeted on improving the ecological risk assessment process. Beyond a simple food chain exposure, other sources of energy and material “loss” must also be accounted for to comprehensively assess effects of stressors on an energetics basis. For example, gaseous products of digestion (gaseous energy [GE]) resulting from microbial activity in the large bowel (and, depending on species, their associated diverticula) may be critical to a complete accounting of energy, particularly if gastrointestinal function is compromised consequent to exposure. Although limited in their contribution to material and energy balance focused on many chemicals of concern, the energy and materials in these gases are unavailable to the animal, as is energy associated with elimination of nitrogenous waste (e.g., ammonia, uric acid, and urea in amphibians and reptiles, depending on species), or urinary energy (UE). Both UE and FE have associated material loss critical to the analysis of a chemical’s disposition in food chain exposures. As a consequence of these losses, DE can be further refined and called metabolizable energy (ME), which is

ME = IE – (FE + GE + UE)

(5.1)

Foods associated with ME yield energy available for metabolism, although part of the energy is inevitably lost to heat production or heat energy (HE). Physiologically, even in ecothermic vertebrates, heat is produced during basal metabolism, digestion, fermentation by enteric microbes, formation of waste products, and other biological processes. Although an animal’s thermoregulatory activities, endothermic, ectothermic, homeothermic, or poikilothermic, are critical to its survival and long-term success as a species, energy lost as heat is not available for processes such as growth, reproduction, foraging, and other metabolicactivities. As such, the net energy (NE) of a biological system is ME less HE, with energy and materials potentially allocated to physiological and biochemical functions such as basal metabolism, maintenance, growth, reproduction, stored and tissue energy production, and detoxification). These physiological and biochemical functions are responsive to various environmental stressors, including chemicals presenting ecological risk. As suggested by Figure 5.1, the flow of material and energy characteristic of a given parcel of food is relatively easy to summarize conceptually. Food chain analyses can be accomplished beyond that captured by simple screening level. The allocation of material and energy can be further quantified within an exposure assessment focused on food chain analysis (e.g. measuring food intake, urinary and fecal outputs, and reproductive parameters) relative to tissue residues and observed adverse effects. However, routes of exposure critical for amphibians and reptiles are not adequately characterized by simple food chain analyses. From an ecological and biological perspective, once a screening

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level analysis of exposure is completed, exposure should no longer simply be a question of knowing a chemical’s concentration in various substrates, foraging items (feed or coincidentally ingested soil or sediment), and a foraging rate. Rather than a regulatory toxicological perspective, an ecological perspective seems more appropriate to the evaluation of chemical risks once the screening level analysis is complete, especially for animals such as reptiles and amphibians whose life histories include critical life stages that are incompletely modeled by a simple food chain analysis.

5.1.2 Physiological Ecology, Exposure Models, and Food Chains Despite their veneer of arithmetic sophistication (see Pilkey and Pilkey-Jarvis 2007 for general discussion of models applied to environmental science), exposure models (Equation 5.2) for terrestrial and semiaquatic vertebrates that focus solely on materials transfer are relatively primitive for reptiles, and even more so for amphibians. E=

1 T

∑C

ijk

tk



(5.2)

where E = exposure concentration or exposed dose, T = total time and space over which the concentrations in various microenvironments or habitats are to be averaged, Cijk = concentration in microenvironment k that is linked to environmental matrix i by pathway j, and tk = time and space that accounts for a receptor’s contact with specific microenvironment or habitat k. Most often, Equation 5.2 is decomposed and simplified, and as a unitless narrative equation yields

ED =

∑ Der

ed

+ Inhed + f(Inhed) + Inged + f(Inged) + DWed + f(DWed)

(5.3)

where ED = exposed dose, Dered = dermal or cutaneous exposed dose, Inhed = inhalation exposed dose, f(Inhed) = exposed dose coincidental to inhalation, Inged = ingestion exposed dose, f(Inged) = exposed dose coincidental to ingestion, DWed = drinking water consumption exposed dose, and f(DWed) = exposed dose coincidental to drinking water consumption. Ultimately, given practical matters and all too frequently, the relatively sparse to nonexistent data available, risk analysts further simplify, which yields a food-chain-dominated exposure model:

ED =

∑ Ing

ed

+ f(Inged) + DWed

(5.4)

where exposed dose contributed by dermal and inhalation routes is considered negligible to the derivation of exposed dose (USEPA 2003, 2005). Subsequently, the fraction of exposed dose that is absorbed is frequently assumed to be 100% or estimated using available literature values to account for efficiencies in uptake of material across surfaces such as the walls of the alimentary canal or other serosal-mucosal barriers between external and internal environments or across cell membranes. Beyond screening level evaluations, these estimates are most often derived in the absence of energy and nutritional considerations for foods in their diet, and the variability linked to, for

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example, seasonal patterns characteristic of an animal’s dietary intake (Robbins 1993; Barboza et al. 2009). Rather, food chain models consider “whole animal” functions of water consumption, foraging rate, food intake, and food processing through lumped exposure parameters (e.g., site use factors, assimilation values). In short, you are what you eat and drink. Within an ecological risk context, food chain modeling has become a primary tool for evaluating exposure in terrestrial and semiaquatic wildlife (USEPA 2003, 2005). Yet depending on the animal at risk and the life history stages most sensitive to exposure (often the early developmental stages), the tool may be relatively ineffective in evaluating potential adverse effects and risks linked to exposures in field settings. For example, discounting dermal exposures and percutaneous uptake to nil is highly problematic, given the natural history and life history characteristics of members of the Amphibia (e.g., their semipermeable skin plays a critical role in respiration, such as cutaneous respiration in the lungless plethodontids, and uptake of materials from the environment and may be a critical route of exposure for some environmental contaminants).

5.2 Pathways of Contaminant Exposure for Amphibians and Reptiles If we adopt a physiological ecologist’s frame of reference in the analysis of exposure and effects, we might appreciate more the value of energetic costs associated with chemical exposure. For example, a more complete accounting of time and energy expended on preying or foraging for food potentially contaminated with environmental chemicals would be gained, and effects linked to reduced nutritive value and altered energetic costs could be characterized. Chemically contaminated food resources may have increased metabolic costs associated with their disposition, for example, detoxification and elimination, or transport to storage compartments. The variety of physiological, morphological, and behavioral adaptations characteristic of biota potentially exposed to environmental chemicals precludes a comprehensive accounting of energetic costs associated with contaminant exposures. Indeed, as Rowe et al. (2003) observed, biological diversity is realized in part because a wide variety of adaptive strategies have evolved toward optimization of reproductive fitness. When biological, ecological, and abiotic conditions in the environment diverge significantly from optimal fitness, population level changes may result. In the following sections, our overview of physiological systems key to exposure will be based on the exposure equation (Equation 5.3) that reflects field settings commonly encountered by amphibians and reptiles. While this field-based exposure equation moves beyond a simple food chain model of exposure and is characterized by sparse data sufficient for analysis, uncertainties in evaluation of ecological risks will only be reduced when life history strategies and an animal’s natural history are considered in total. Inconvenience should not foster complacency in the risk assessment community. Rather, developing models based upon numerous life history traits can aid in identifying research that will benefit the screening level risk assessment process, while enhancing our understanding of the herpetofauna and their adaptions to the environment.

5.3 Physiological Ecology of Amphibians and Reptiles: Natural History and Life History Attributes Influencing Exposure Exposure of herpetofauna occupying aquatic and terrestrial habitats is no less significant than that for other wildlife. As McDiarmid and Mitchell (2000) observed, an animal’s daily, seasonal, and annual movements undoubtedly affect exposure in the herpetofauna, given the relatively long distances some species move. Indeed, exposures may be as complex as any envisioned for other species at risk, and for many herpetofauna, these movements may encompass a wide range of

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aquatic, wetland, and terrestrial habitats in their sojourns across the landscape. Movement across heterogeneous habitats will likely be overlain with a similarly heterogeneous range of types and concentrations of environmental chemicals and other stressors, which will complicate forecasting models. For example, in typical exposure models bioaccumulation is often dominated by dietary exposures (food and water consumption) in terrestrial vertebrates, yet chemicals occurring in the exposure matrix may be readily absorbed across dermal epithelia or volatilized from a solid or liquid phase into air, providing other means of exposure (see, for example, Noble 1931; Boutilier et al. 1992; Shoemaker et al. 1992; Duellman and Trueb 1994; and Stebbins and Cohen 1995 regarding cutaneous respiratory surfaces, which may serve as potential routes of uptake). Many variables affect the magnitude of bioaccumulation in terrestrial exposures in adults or early developmental stages (e.g., in reptile eggs). While the transfer of chemicals within food chains may conveniently be described by transfer coefficients or functions that characterize the relationships among trophic levels (Pastorok et al. 1996; Pascoe et al. 1996; Linder et al. 1998, Linder and Joermann 1999), we have only a limited understanding of the interrelationships among other exposure matrices and critical life stages in the ontogeny of amphibians and reptiles. These environmental factors may be abiotic, such as physicochemical characteristics of chemical or exposure matrix (sediment or soil), or biological in character, such as life-history-dependent attributes related to gastrointestinal or nutritional physiology, foraging, or food preference (see Larsen 1992; Hamelink et al. 1994; Langston and Spence 1995; Linder et al. 2002). Needless to say, beyond the convenience of screening level exposure models, the perspective of a physiological ecologist would undoubtedly contribute to developing improved models by investing research into the basic physiology of the herpetofauna. The following sections will focus on the physiology of amphibians and reptiles that bear directly on exposure models such as Equations 5.3 and 5.4 that are called on by risk analysts as part of their evaluation of ecological risks.

5.3.1 Dietary Exposures, Gastrointestinal, and Digestive Physiology The significance of gastrointestinal and digestive physiology in the evaluation of dietary exposures to environmental chemicals has received little, if any, study by ecotoxicologists working with herpetofauna. Physiologists of various persuasions, however, have devoted much effort toward characterizing the anatomical, morphological, and physiological responses of the vertebrate gut to a wide range of food sources and environmental stressors (see, e.g., Secor 2005a, 2005b). Indeed, researchers and regulators would benefit from an increased knowledge of these organismal responses to the range of stressors that inevitably affect responses of herpetofauna to environmental chemicals in the field. For example, Secor (2005b, p 282) observed that “vertebrate intestinal tracts possess an array of structural and functional adaptations to the wide diversity of food and feeding habits,” which inevitably begs the question of what role this variability plays in evaluating dietary exposures to environmental chemicals in herpetofauna. Furthermore, anatomical, morphological, and physiological differences associated with diet (e.g., differences between alimentary canals of herbivores and carnivores), and the adaptive plasticity of the gut clearly point to sources of uncertainty that currently are not captured by regulatory applications common to the risk assessment process. The capacity to which intestinal performance responds to changes in digestive demands is a product of evolutionary and cellular mechanisms (Secor 2005b). Inevitably, exposure of herpetofauna to chemical stressors in the field cannot help but be complicated by feedbacks that control function and structure of the gut in unchallenged systems. The issues associated with this dietary route of exposure become even more prominent when static “snapshots” focused on threshold values, such as toxicity reference values (TRVs), are applied to the ecological risk assessment process. While consensus TRVs for amphibians and reptiles are lacking, the paucity of data encourages that their future development be by design rather than limited by the convenience of existing data. Indeed, life history attributes of

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the herpetofauna that influence exposure must not be undervalued in developing threshold values. The current lack of TRVs for amphibians and reptiles gives us time to develop regulatory benchmarks that are ecological in nature and more adequately estimate risks to herpetofauna. Uptake of environmental chemicals by herpetofauna resulting directly from dietary exposures may be simply viewed as suggested in exposure models currently in vogue. Yet, beyond a screening level analysis, the simple exposure model exemplified in Equation 5.3 or 5.4 undervalues our current understanding of some dietary constituents such as metals and some organics that have nutritive roles, or for those whose uptake is mediated by gastrointestinal processes characteristic of the herpetofauna. For example, the regulation of intestinal performance (i.e., the total small intestinal capacity to absorb nutrients estimated as a product of small intestinal mass and mass-specific rates of nutrient uptake [Secor 2005a]) varies with type of food consumed and level of feeding activity. From a physiological ecologist’s perspective, intestinal performance may be conveniently categorized, based in part on life history attributes that capture the range of exposures potentially associated with amphibians and reptiles. It is not a “fits-all-sizes” world when dietary routes of exposure are considered for the herpetofauna, particularly within the context of exposure to chemical stressors in the field. For example, “sit-and-wait foragers” (e.g., common to some snakes) rely on ingesting a single, frequently large meal followed by an extended nonforaging, resting state. This life history strategy is also represented by other estivating or hibernating herpetofauna, which experience long episodes of fasting accompanied by downregulation of intestinal morphology and function. Not surprisingly, fasting reduces energy expenditure during these extended periods. In contrast to these sit-and-wait foragers, “frequently feeding foragers” regulate intestinal performance by alternately fasting and feeding to earn energy savings that offset costs associated with upregulating the gut during feeding episodes. In the herpetofauna the regulation of intestinal performance varies widely, in part as a function of the degree to which mass-specific rates of nutrient transport are depressed due to loss of intestinal mass during fasts common to these vertebrates. While models focused on dietary routes of exposure may be simple implementations of food chain analysis, these analyses for wildlife do little beyond providing a simple, screening level effort potentially disconnected from reality. Simply stated, the physiology of the intestinal tract drives the relationship between consumption and assimilation. The diversity of vertebrate food and feeding habits is matched by an array of adaptive intestinal morphologies and physiologies (Stevens and Hume 1995; Karasov and Hume 1997), which potentially could be incorporated into refined exposure models. For example, compared to carnivores, the intestinal tracts of herbivores are longer, more complex, and generally include fermentation chambers (e.g., intestinal diverticula). The longer herbivore gut is necessary because plant material is more difficult to digest than animal material (Stevens and Hume 1995). Physiologically and biochemically, intestinal tracts of herbivores hydrolyze and transport simple sugars at greater rates relative to amino acids. In contrast, the intestinal tracts of carnivores generally favor transport of amino acids (Karasov and Diamond 1988; Stevens and Hume 1995). Differences in substrate breakdown and transport between herbivores and carnivores need to be considered when developing more comprehensive and realistic models of exposure, particularly in relationship to the effects and the metabolic fate of the chemical once ingested. Among the many species of vertebrate, the gut — physiologically and morphologically — is highly malleable and adaptive to changes in digestive demand. Environmental factors such as seasonal changes in feeding regimens (primarily, changes in diet or in feeding behaviors) influence these physiological and morphological changes and their associated changes in energy requirements and ingested nutrients (Piersma and Lindström 1997). The most frequently noted responses to changes in digestive demand are morphological, for example, increased or decreased intestinal mass (Karasov and Diamond 1983) and physiological, as exemplified by altered modulation of intestinal function (e.g., changes in enzyme activity and nutrient dynamics; see Karasov and Hume 1997). Not surprisingly, much of the work characterizing gastrointestinal responses to environmental stressors and changes in nutritional physiology and ecology was initially completed in birds

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and mammals, but regulation of intestinal performance among amphibians and reptiles has gained increased research interest and has been recently reviewed (see, e.g., Secor 2005a, 2005b). The gastrointestinal (GI) tract, like other major organ systems, can experience a wide range of demands. However, unlike renal, pulmonary, or cardiovascular systems, which must provide at least a minimum level of performance under nominal environmental conditions, GI tracts of many organisms are routinely quiescent during periods in an animal’s lifespan. The duration of quiescence will vary with species, ranging from little, if any (e.g., common to endotherms classically considered as having high metabolic demands), to extended fasts with associated gut quiescence ranging from a few hours (e.g., common to organisms having daily feeding patterns) to several months (e.g., commonly associated with periods of hibernation, estivation). As a shared attribute throughout the vertebrates, feeding or foraging strategies of the herpetofauna vary widely, depending on species and habitat. In some vertebrates, extended, nearly yearlong fasts may be a common species attribute. For example, female rattlesnakes (Crotalus viridis) may not feed during breeding and perinatal periods, which means individuals will fast for 2 hibernating cycles and the intervening summer (Macartney and Gregory 1988). Amphibians and reptiles that feed frequently generally possess a relatively limited range in digestive performance with neither feeding nor fasting occurring as dominant features in their natural history. At the other end of the spectrum, those animals having life histories characterized by long episodes of fasting tend to regulate digestive performance much more widely with feeding and fasting. As Secor (2001) observed, these bounding patterns of digestive regulation are primarily distinguished by the ability to “upregulate” or “downregulate” digestive performance, depending on species characteristics and food availability. Downregulation of digestive performance in ecotherms such as amphibians and reptiles is linked with fasting that commonly occurs as aphagia when the gut is quiescent (Gregory 1982; Pinder et al. 1992). During these fasts, animals depend on stored energy to meet metabolic needs, and adaptive responses reduce energy expenditures to ensure survival (see Secor 2005a, 2005b). Depending on species, differential capacity to regulate digestive performance will be linked to time-dependent factors that influence the gut’s response, such as duration of the fast, which inevitably influences chemical exposure through the diet. The wide range of life histories characteristic of amphibians and reptiles captures a similarly wide range of feeding habits and digestive performances. Secor (2005a, 2005b) noted that a wide regulation of intestinal performance is exhibited within the herpetofauna, where fasts may occur for months and yield 5- to 30-fold changes in nutrient uptake capacity. Frequently feeding species having limited fasting periods vary with respect to nutrient uptake capacity; many display only slight increases, while others (e.g., Ambystoma tigrinum larva, Bufo speciosus, and Rana pipiens) present significant increases in nutrient uptake capacity with feeding. Even a cursory review of Stevens (1988) or Stevens and Hume (1995, 2004) clearly suggests potential research opportunities for physiological ecologists who expand their research horizons to include ecotoxicological issues related to exposure modeling. The vertebrate gut, then, potentially displays variability in its morphology and physiology as a function of digestive demand, and in its function within the context of natural history strategy (e.g., feeding habits and food preferences). The capacity to regulate digestive performance in response to variation in digestive demand is a common adaptation associated with a range of adaptive strategies, and generally reflects a product of changes in size, type, or an interaction of meal size and type. Developmental changes (e.g., larvae to adult) are also commonly observed pressures that influence the function and structure of the vertebrate gut, which necessarily affects exposure to environmental chemicals as constituents in the diet. Adaptive strategies seen across the spectrum of vertebrate life histories may be shaped relatively little in those species displaying little variability in meal quality or quantity, but in those species having a widely varying diet, digestive capacity may be similarly variable, which dramatically influences adaptive responses observed in gut structure and function. Digestion and nutrient uptake involves components of the gut working in concert with other visceral organs in an integrative response initiated by the feeding event, an event that has become a dominant focus in contaminant exposure modeling. An integrative approach incorporating chemical

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stressor exposure and uptake of nutrients could not help but improve our understanding relationships between effects and “exposed dose–chemical assimilation–absorbed dose.” Interactions potentially in play for the regulation of tissue performance vary with digestive demands, which likely means the greatest metabolic and functional demands on an organism occur during digestion or during other highly integrated functions, such as reproduction. Hence, maximum digestive demand may set the upper limits to the performance of supportive physiological systems considered next.

5.3.2 Thermoregulatory, Osmoregulatory, and Excretory Physiology Within the context of exposures in the field, an initial consideration of whole animal responses to environmental factors focused on temperature and water is critical to any stressor evaluation. Physiologically, amphibians and reptiles are similar, yet sufficiently different to warrant a brief overview of their physiological attributes related to interrelationships in functions managing temperature and water homeostasis. Amphibians and reptiles are ectothermic; hence, ambient temperatures affect molecular and cellular processes (e.g., enzyme activities and protein synthesis), which are commonly observed as integrated organ system responses in the animal, such as digestion, and whole-animal regulatory functions related to sensory and behavioral interactions between organism and the environment (Wieser 1973). In amphibians, for example, temperature coupled with humidity will influence reproductive activities such as emergence (Bellis 1962; Cree 1989), vocalization, egg deposition (Blair 1961), embryonic development (Herreid and Kinney 1967), growth (Bellis 1962; Berven et al. 1979), metamorphosis, and the immune response (Maniero and Carey 1997; Jozkowicz and Plytyez 1998; Taylor 1998). For amphibians, temperature and moisture affect physiological and behavioral responses to contaminant exposure, particularly given these factors’ roles in up- or downregulation of metabolic processes that influence chemical uptake, metabolite production, and clearance from the system. Within species and among populations, thermal limits vary geographically, seasonally, and diurnally. An individual’s previous experience with temperature extremes will also influence exposure (Berven 1982). Thermal tolerance and preferences can be altered by exposure to chemicals such as organophosphorus compounds (Johnson 1976; Johnson and Prine 1976) and other chemicals that could interfere with endocrine and neuroendocrine regulation (Hutchinson and Dupre 1992). Many amphibian life history attributes and environmental cues keyed to homeostatic adaptations and attributed to temperature and thermoregulation are ultimately linked to osmoregulation and water conservation or repiratory gas exchange (Bellis 1962). Amphibians generally tolerate temperatures below their preferred temperature better than above their preferred range of temperatures. Temperatures greater than their preferred upper limit are linked to excessive water loss and are accompanied by increased risks of desiccation. Short-term responses to increased temperature depend on the duration of the thermal or osmotic stress, and may elicit integrated responses to that challenge. For example, amphibians may respond physiologically and modify behaviors in response to temperature challenge (Londos and Brooks 1990; Rome et al. 1992). Behavioral changes include basking, adjusting their posture to minimize or maximize surface-to-volume ratio for both thermal and water regulation, and aggregating to areas in which temperatures may be cooler, more stable, and less affected by solar radiation fluctuations (e.g., under forest substratum or at the bottom of ponds). Prolonged physiological adjustments include developmental adaptations, modifications in ventilation, metabolism, glomerular filtration, and hormonal feedback (Kim et al. 1998). During seasonal or otherwise prolonged changes in environmental temperature and humidity, amphibians can become dormant and enter hibernation or estivation. In the dormant animal, nonessential behaviors are reduced and metabolism is lowered to minimize depletion of energy stores and maximize survival. Hibernation and estivation are adaptations to environmental stressors such as temperature extremes or reduced food resources. If resources are not sufficient to maintain their lower bounds of metabolic activity required for day-to-day activities, attendant outcomes are starvation and energy

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depletion (e.g., Scott 1994). Hibernation is a common response to the cold winter of temperate climates and continues to be a research topic of keen interest to physiological ecologists. Indeed, hibernation represents a topic that could warrant review by ecotoxicologists, since the interrelationships between temperature and reduced water resources would likely influence exposures to chemical stressors in field settings. Temperate amphibians, ranging from those dominantly aquatic in their habitat requirements to those strongly terrestrial in their preferred habitats, overwinter in hibernacula that range from below-frost-line burrows in soil to shallow digs in near-surface materials (e.g., forest duff) to sediments in aquatic habitats. For example, many temperate frogs overwinter buried to varying depths in surface litter and soils, where the animals are likely to be exposed to dehydrating conditions and subzero temperatures, which they can survive by virtue of their profound tolerances to dehydration (Hillyard 1999) and freezing (Schmid 1982). Wood frogs (Rana sylvatica) range further north than other anurans, and recover from severe dehydration (Churchill and Storey 1993) and survive the freezing of up to 70% of their body water at temperatures between –4 and –6 °C (Storey and Storey 2004). Various molecular, biochemical, and physiological adaptations provide freeze-thaw tolerance of tissues, since preventing deep freeze and internal fluid crystallization is critical for survival (see Tattersall and Ultsch 2008). Estivation occurs during prolonged heat or drought conditions. Desert spadefoot toads (Scaphiopus couchii and S. hammondii), for example, inhabit areas in which it may not rain for up to 1 year. As in hibernation, gluconeogenesis is initiated prior to estivation to accumulate energy reserves, then metabolism is reduced, although aerobic respiration is maintained. If the animal becomes too dehydrated, ventilation rate is lowered, increasing the risk that toxic levels of CO2 will accumulate in the blood (Pinder et al. 1992). Examples of estivating behaviors to decrease water loss include burrowing further into moist, cool soils and forming cocoons. Cocoons can be formed by wrapping a layer of mud obtained from the bottom of a pond or by shedding dead epidermal layers of squamous epithelial cells to completely encase the estivating individual. The oral cavity is left open and pulmonary gas exchange predominates, effectively reducing water loss by 90 to 95%. Conserving water is the prominent determinant contributing to onset of hibernation or estivation. Again, depending on the habitats occupied during estivation, environmental chemicals may contribute to the multiple stressor exposure experienced by the animal. In amphibians, most water, ion, and gas exchange occurs through the skin; however, dermal uptake and osmotic regulation requirements differ between aquatic and terrestrial systems. Hence, dermal exposure and consequently response to chemicals may be different (Cree 1989). For example, larval epidermis in amphibians is ciliated and composed of several cell layers overlying a thin basement membrane (Burggren and Just 1992; Duellman and Trueb 1994). Throughout development, the dermis and epidermis are restructured, thickened, or keratinized; dermal glands are developed and pigmentation is changed. Adjustments in skin permeability are under neural and hormonal control (Reboreda et al. 1991). Whereas drought conditions increase water uptake and could place amphibians at a higher risk to waterborne chemicals, adaptations such as cocoons may protect individuals temporarily from direct exposure to chemicals in soil, water, or air (Stiffler 1988). Some amphibians may be at greater risk of chemical exposure during the early breeding season. Osmotic permeability fluctuates seasonally, accounting for the spring water drive during which terrestrial amphibians migrate to the breeding ponds (Duvall and Norris 1980; Boutilier et al. 1992). Plethodontid adults that rely more than 90% on transepithelial respiration may also be at higher risk of exposure. Water imbition occurs through percutaneous routes in amphibians and is primarily mediated through epithelia that have a long history of studies focused on ion transport and osmotic water uptake from the organism’s environs (Jørgensen 1997; Larsen 1991). Uptake of water and ionic solutes is followed by internal processing, with final disposition occurring, in part, through the excretory system. Amphibians are relatively intolerant of salt challenge; hence, most amphibians are limited to freshwater environments. To maintain homeostasis, Na+ ions and K+ are transported into the intestines, where Cl– efflux is coupled to Na+ influx. Na-K ATPase is present throughout extracellular compartments. As with many cellular processes, changes in environmental pH affect ion

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transport (e.g., acidic pH inhibits Na+ uptake and jeopardizes the integrity of the outer membrane’s tight junctions, yielding intracellular ionic depletion; Freda and Dunson 1984; Freda 1986). Ionic regulatory systems involving Na+ and acid-base coupling also are active in the amphibian renal and bladder systems (Rohrbach and Stiffler 1987; Stiffler 1988; Toews and Stiffler 1989). Despite slight differences in the rate of glomerular filtration between salamanders and anurans, the kidneys of both groups produce large volumes of dilute urine when the animal is in freshwater. When water is unavailable, filtration rates drop and no urine is produced; hence, water is conserved. The bladder and lymph sacs continue to resorb and store water during dehydration, but the tight epithelium becomes selectively permeable to ions and water, so that urine formation and glomerular filtration can be controlled (Tufts and Toews 1985). Frogs can excrete urate salts to further rid the dehydrated system of excessive Na+ and K+. Depending on environmental conditions, amphibians have proven adaptive to a variety of dehydrating habitats. For example, spadefoot toads store urea in plasma and muscle as a mechanism to offset hydrostatic pressure, and amphibians living in brackish waters (e.g., Rana cancrivora and Bufo marinus) can retain protein by-products (e.g., urea and amino acids) to maintain osmotic equilibrium. Unlike amphibians, many reptiles maintain body temperatures as high as or higher than those of endotherms (birds and mammals). However, in contrast to birds and mammals that rely on metabolic heat to maintain homeothermy, reptiles generally rely on external heat sources. While some sea turtles and pythons generate significant metabolic heat, most reptiles are poorly insulated by prominent layers of subcutaneous fat (except for leatherback sea turtles [Dermochelys coriacea]), fur, or feathers. Without insulation, metabolic heat is quickly lost. As do poikilotherms in general, reptiles characteristically display body temperatures that fluctuate rather than being controlled by physiological means. Yet, reptiles do exhibit a range of thermoregulatory ability, and control of heat gain and loss from external sources and maintenance of body temperature are achieved primarily through behavioral mechansisms. Thermoconformers regulate their body temperature to a limited extent, and generally track environmental temperatures in the classic definition of a poikilotherm. Such species generally live in relatively invariant thermal environments such as those living in water, caves, or burrows or under thick forest canopy. On the other hand, thermoregulators control their temperature very precisely — they make for a very good homeotherm or, more specifically, an ecothermic homeotherm — often through a combination of behavioral and physiological means. Body temperature is regulated to optimize physiological processes, and if environmental conditions are sufficient, the range in which a reptile regulates its body temperature is commonly the species’ activity-temperature range. Within the activity-temperature range is the selected (preferred) temperature range, which is more narrowly regulated if possible. Behavior is critical to thermoregulation in reptiles, and without access to a range of environmental temperatures or heat sources, most reptiles will have a body temperature close to ambient. However, if variations in environmental temperatures do exist in their habitat, many reptiles can significantly regulate their own body temperature. This behavioral temperature regulation, especially in large-bodied species, can result in a near homeothermic body temperature. As a heat transfer medium, blood and its flow within vascular networks is key to regulating body temperature. Reptiles have a variety of anatomical and physiological means to alter their thermal conductivity by varying blood flow, including cardiac and vascular shunts, altered heart rate, vasoconstriction, vasodilation, and countercurrent heat exchange mechanisms (Espinoza and Tracy 1997). Adaptations for water conservation and osmoregulation differ between amphibians and reptiles. Reptiles have several major adaptations for terrestrial life, such as more intricately structured, yet less permeable skin, a more advanced urinary system, and shelled amniotic eggs. These adaptations permit reptiles to survive in terrestrial habitats in the absence of sources of freshwater critical to the reproductive biology of nearly all amphibians. Although 1 species of turtle (northern long-necked turtle, Chelodina rugosa) lays eggs underwater in sediments or submerged soils (Kennett et al. 1998), reptiles generally rely on terrestrial habitats for egg laying. Even oviparous aquatic species such as sea turtles return to land to lay their eggs.

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Structurally, reptilian skin is a relatively complex, layered organ in contrast to the skin of amphibians. In reptiles the upper layers of skin are highly keratinized, with keratin being produced by keratinocytes in the basal layers (stratum germinativum) of the skin. Keratin represents a major proportion of the skin’s proteins in the form of scales. Although the keratin helps reduce water loss and protects the skin, it is not completely impermeable to water. The major barrier to water imbition is a layer of lipids within the skin, which varies significantly among species. Percutaneous and dermal routes of exposure should not be dismissed as insignificant, especially in those species that have a reduced underlying lipid layer that, in its absence or reduced state, may increase likelihood of uptake of polar, waterborne compounds. Alternately, those with relatively impermeable skin, due to cutaneous layers of lipid, may be more prone to cutaneous absorption of lipophilic compounds. From a comparative perspective, the excretory system of amphibians and reptiles represents a snapshot in the evolution of the kidney as an osmoregulatory organ. In adult amphibians, the kidney is a modified mesonephros and includes development of some features characteristic of the metanephric kidney, for example, the absence of nephrostomes and the joining of nephric tubules to collecting ducts that have developed as outgrowths of the mesonephric ducts that occur in the kidney’s metanephric zone (Clothier et al. 1978; Meseguer et al. 1978; Hinton et al. 1982; Sakai and Kawahara 1983; Uchiyama et al. 1990; Møbjerg et al. 1998). Due to its unique structure and associated function, the amphibian kidney is often regarded as the opisthonephros (Kardong 2005). In contrast, the mesonephros in reptiles forms the functional excretory organ through early development (in ovo); then at hatching, the metanephric kidney supplants that embryonic structure in excretory functions and water-solute conservation. Although all reptiles possess metanephric kidneys, they are highly variable in size and shape. For example, in turtles and most lizards, the urinary bladder arises from the ventral wall of the cloaca and is connected to the kidneys via the ureters. In contrast, a urinary bladder is absent in snakes and crocodilians, and the kidneys empty directly into the cloaca (Zug et al. 2001). The disposition of nitrogenous wastes differs between the classes of herpetofauna. From an energetics perspective, ammonia is relatively inexpensive to metabolize, but is highly toxic in all but the most dilute solution. Hence, aquatic organisms have a relatively great advantage in handling nitrogenous wastes as ammonia, and avoid ammonia toxicity by its dissolution in large quantities of water and its excretion in dilute solution. In adult amphibians and reptiles, ammonia is commonly converted into urea or uric acid, as an adaptation responsive to ammonia’s high toxicity in wetland and terrestrial environments. At the same time, both urea and uric acid require less water in their synthesis; hence, water is conserved, which is also a driving force behind adaptations to habitats characteristic of many herpetofauna. As with ammonia, urea is highly soluble in water, but it is much less toxic. Urea can also be accumulated to relatively greater concentrations in tissues without undo adverse physiological effects and can be excreted in a concentrated form. Although relatively costly to metabolize, the synthesis of urea from ammonia and carbon dioxide (Grisolia et al. 1976) benefits the water conservation processes necessary for survival in terrestrial habitats. In contrast to ammonia and urea, uric acid is nearly insoluble in water, but nonetheless, is primarily excreted by the kidney. In reptiles, uric acid passes to the cloaca via the ureters. Water is highly conserved in the disposition of nitrogenous wastes as uric acid, but incurs high energetic cost in its production. Most adult amphibians are ureotelic, although larval tadpoles or highly aquatic adults rely to varying degrees on ammonia as the dominant form of nitrogenous waste. Reptiles tend toward reliance on uric acid as their nitrogenous waste, although the class is far from exclusively uricotelic. In reptiles inhabiting freshwater habitats (e.g., crocodilians), elimination of nitrogenous wastes while maintaining osmotic balance is not a problem, since the influx of water into the body and the osmotic loss of ions occurs as a consequence of their environs. The flux of water and ions is reduced in part by the keratinized skin of reptiles, and the kidneys produce dilute urine; hence, ions are conserved, water in excess is eliminated, and nitrogenous wastes occur as ammonia (Minnich 1982). In terrestrial and marine environments, additional osmotic stress is experienced by reptiles.

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Although some terrestrial reptiles still eliminate a significant proportion of their nitrogenous wastes as ammonia, others reduce urinary output to conserve water and concentrate nitrogenous wastes as either urea or uric acid (Minnich 1982). Tortoises and turtles occupying terrestrial habitats tend to be uricotelic, as do some anurans. Marine environs pose similar challenges with the added hazard of an influx of salts. Because their metanephric kidneys cannot produce hypertonic urine, marine reptiles have adapted various glands for elimination of excess salt. For example, in estuarine crocodiles, special glands on the tongue excrete salt (Taplin and Grigg 1981), and in marine iguanas, nasal glands produce concentrated brine that is expelled from the nose (Dunson 1969). Similarly in sea turtles, lacrimal glands produce a constant efflux of salt (Schmidt-Nielsen and Fänge 1958). The driving force for metabolic adaptation and the nitrogenous wastes characteristic of a species are osmotic stresses placed on the reptile by its environment. In part, these stresses reflect inputs from various routes of exposure, including dermal and percutaneous pathways, and those linked with respiration and gas exchange.

5.3.3 Dermal and Percutaneous Exposure, Respiratory Physiology, and Gas Exchange From the perspective of exposures to multiple environmental stressors — be they anthropogenic or natural in origin — integrated responses are likely due to functional overlaps among the broadly characterized physiological categories considered in Sections 5.2.1.1 and 5.2.1.2. Simply stated, exposures dominated by dietary routes may be mediated by gastrointestinal and digestive systems of exposed biota, yet those responses may inevitably be tied to responses linked to thermoregulatory, osmoregulatory, or excretory functions. Similarly, interrelationships between functions considered under the auspices of thermoregulatory, osmoregulatory, and excretory physiology, and those linked to respiratory physiology and gas exchange share a route of exposure reflected in exposure models such as that presented in Equation 5.3, wherein dermal and percutaneous routes of exposure variously contribute to total exposed dose realized in field settings. Mechanisms of gas exchange reflect differences in selective pressures characteristic of aquatic and terrestrial environments. In aquatic habitats, oxygen is less available and more difficult to extract from water, while in terrestrial systems, oxygen is readily available for transcutaneous diffusion across respiratory membranes. Carbon dioxide, however, is not as readily released to the atmosphere unless respiratory systems are functionally and structurally adapted to promote release of CO2. Hence, respiratory systems in amphibians vary with life stage; for example, the primary respiratory organ in many amphibian embryos and larvae is the epithelia of the skin (Seymour and Bradford 1995). Within each life stage, species-specific functional and structural adaptations to the environment reflect the wide range of responses associated with constraints on gas exchange in aquatic and terrestrial habitats (e.g., the presence or absence of ventilating systems such as lungs, availability of a surface for diffusion and gas exchange, and the concentration of oxygen in the surrounding medium). For amphibians, life stage and abiotic and biotic characteristics of the environment determine the mode and effectiveness of gas exchange. Within the range of environmental conditions that limit an animal’s distribution across the landscape, respiratory functions provide means that ensure gas exchange and contribute to acid/base homeostasis within the animal. Feedback systems serve to control exposures to environmental stressors, such as reduced dissolved oxygen in surface water, and responses to these exposures that are potentially linked to adverse effects (e.g., hypoxia or hypercapnia). At the same time, these respiratory responses, and accompanying morphological and behavioral changes potentially linked to these physiological responses, likely influence exposures to environmental chemicals. Although the primary site for gas exchange in amphibians is the skin, differences among species and life stages reflect conditions in their environment. These differences in part are due to morphological adaptations for gas exchange throughout the range of habitats where the herpetofauna occur. For example, in the aquatic larvae of amphibians, branchial uptake is linked to the ability to filter feed through the use of a buccal pump mechanism. Filtered water entering the mouth and nostrils is

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pumped through a pharyngeal cavity over gills into an opercular chamber and out through spiracles. Aquatic salamanders frequently display neotenic traits, most notably the retention of external gills that allow water entering the mouth to pass over gill slits for gas exchange. To some extent, gill growth is oxygen dependent, with elevated O2/CO2 ratios suppressing growth and elevated CO2/O2 ratios promoting growth and branching (in Ambystoma maculatum; see Branch and Taylor 1977; Duellman and Trueb 1994). Respiration in amphibians that inhabit well-oxygenated mountain streams is primarily transdermal; therefore, gills and lungs (if present) are reduced. Gills in species occupying lentic habitats are more developed and have increased surface area on the villi for oxygen uptake. Amphibians inhabiting lentic conditions may also undulate in the water to further increase convection and gaseous exchange. Under anoxic conditions, gill ventilation ceases and transepithelial loss of oxygen predominates. Aquatic animals can reduce oxygen demand either by decreasing activity (thereby increasing the risk of anaerobic accumulation of lactic acid) or by rising to the water’s surface to gulp air (thereby increasing risks of detection by predators; Kramer 1988; Boutilier et al. 1992). Anurans and salamanders with developed lungs typically respond to oxygen stress by swimming up to the water’s surface to gulp air, yet this characteristic stress response would not be observed in some Bufo spp. tadpoles that lack lungs. However, Bufo spp. will swim to the surface of the water to absorb oxygen through cutaneous respiration (Duellman and Trueb 1994). Gas exchange through the gills is diffusion rate limited, but transcutaneous transfer can be up- or downregulated to some extent by changes in dermal folding and glandular/mucous secretions. Gills also have a role in ion exchange; hence, species-specific anatomy and physiology of gills may be important in explaining differential sensitivity to some chemical exposure (Honrubia et al. 1993; Lajmanovich et al. 1998). When larval gills are resorbed, gill slits are closed and lungs develop. In urodeles, specialized vessels and noradrenergic inputs shunt blood from the gills to the lungs (Malvin 1989). With the exception of both neotenic salamanders, which retain their gills, and members of the lungless Plethodontidae, adult caecilians, salamanders, and anurans use transepithelial, pulmonary, and buccopharyngeal respiration (see Duellman and Trueb 1994). In Plethodontidae, 90% of respiration is dermal, and 10% of the total respiratory vasculature is in the epithelial lining of the mouth. The capacity for diffusion across pulmonary tissue is greater than that of the skin surface, but under conditions commonly encountered in humid, terrestrial habitats, transepithelial diffusion of oxygen from very dense air into the tissues is the predominant respiratory process. Effective ventilation in adults reliant on respiratory surfaces of the lung will depend on lung characteristics such as pulmonary surface area, tidal volume, ventilation rate, capillary density, diffusion distance, and partial pressures. Amphibian lung volumes vary with environmental conditions. Among coldwater inhabitants, lung volumes are relatively small when compared to the more substantial structures characteristic of terrestrial and some neotenic salamanders. The size of the lung also reflects environmental conditions. For example, in Rana catesbeiana lung size increases in response to hypoxia; arterial partial pressure of oxygen, capillary distribution, and rate of blood flow can also be altered in response to changes in surrounding gases (Burggren and Mwalukoma 1983; Duellman and Trueb 1994). Chemoreceptors in the brain alter ventilation rates and other behaviors in response to decreased environmental O2, or increased oxygen demands, and to elevated CO2 concentrations (Shoemaker et al. 1992). During hypercapnia, some species such as adult Bufo spp. increase pulmonary ventilation to reduce CO2 and acidosis. In contrast, aquatic amphibians actively transport HCO3 and H+ and Ambystoma spp. to increase ion excretion. Metabolic or respiratory acidosis linked to increased activity can stimulate the renin-angiotensin system, increasing aldosterone circulation to promote cutaneous excretion of acid equivalents (Eskandari and Stiffler 1997). Dehydration, on the other hand, tends to decrease transepithelial oxygen uptake to reduce water loss that would be experienced with increased ventilation. In such cases, ambystomids compensate by moving to more hydric soils and by increasing glandular secretions that help in water conservation, while enhancing gaseous exchange at the skin surface.

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Efficient oxygen distribution throughout the body is dependent on blood flow, hematocrit, and blood-oxygen carrying capacity (Taketa and Nickerson 1973a, 1973b; Boutilier et al. 1992). Erythrocytes are nucleated in all known amphibians except Plethodontidae. Aquatic species typically have higher blood volumes than do their terrestrial counterparts and also have elevated oxygen carrying capacities, which help in diving and during periods of prolonged submersion (Boutilier and Shelton 1986; Burggren and Just 1992). As a class, reptiles occupy a variety of habitats that have driven development of a wide range of adaptations to ensure adequate gas exchange. Because reptiles occupy habitats ranging from fully aquatic to completely terrestrial, a variety of morphological and anatomic adaptations have evolved that enable body surfaces as effective gas exchange surfaces. In some reptiles, multiple respiratory surfaces may be used at any particular time, depending on habitat or fluctuating environmental conditions. Throughout the reptiles, lungs are the principal respiratory organs for adults. Structually, the reptilian lung is markedly underdeveloped in comparison to birds and mammals. Their lungs have bronchi connected to a relatively simple, sac-like strucuture having a limited number of alveoli for increasing surface area. The alveoli are vascularized and serve for gas exchange. Some reptiles (e.g., varanids, crocodilians, turtles, agamids, iguanids, and chameleons) have a unicameral lung in which the sac-like center of the lung is surrounded by a few septa (Bickler et al. 1985; Powell and Gray 1989). The chambers formed by these septa are each supplied by a bronchiole and are partitioned further by alveoli. Because of their elongated body form, snakes have special adaptations in their respiratory structures, and typically have an enlarged right lung and a rudimentary left lung. The elongated lungs of snakes may be divided into an anterior bronchial lung and a posterior air sac (Stinner 1982). Many snakes also possess a tracheal lung, which is structurally similar to the functional lung but branches from the trachea from where the tracheal rings are incomplete dorsally. The mechanics of inhalation and exhalation are accomplished through a variety of adaptations. Air is tidally exchanged through the lungs by thoracic aspiration. While crocodilians possess a diaphragm that contracts for inhalation and abdominal muscles that contract for exhalation, most reptiles achieve inhalation by contraction of the intercostal muscles; exhalation is passive, resulting from elastic recoil of the thoracic cavity and the weight of the internal organs upon the lungs. Turtles present unique adaptations for respiration because the thoracic cavity is enclosed within a rigid shell. In turtles, the posterior abdominal muscles and pectoral girdle muscles can expand the body for inhalation. Exhalation is accomplished either passively or by retraction of the limbs into the shell, compressing the internal organs and lungs. Other structures also can provide important surfaces for gas exchange. Oxygen and CO2 exchange can take place across the skin, buccopharyngeal, or cloacal surfaces (Dunson 1960; Girgis 1961). To increase gas diffusion, filamentous projections occur on the pharyngeal surfaces of some softshell turtles (Girgis 1961; Winokur 1973). More than a single mechanism may be used for respiration at any time, and methods can change, depending upon circumstances. Although aquatic turtles normally breathe air, during prolonged submergence such as hibernation, buccopharyngeal respiration may dominate gas exchange. Softshell turtles, for example, can obtain as much as 50% of their oxygen requirements from cutaneous and buccopharyngeal respiration (Zug et al. 2001). Similarly, Sabenau and Vietti (2003) considered respiratory and metabolic acidosis and the importance of recovery periods to loggerhead turtle (Caretta caretta) confronted by repeated submergence challenge. In turtles, as well as other diving semiaquatic vertebrates, recovery from any physiological acid-base disturbance is accomplished, in part, by immediately surfacing after the dive, hyperventilating, and resumption of normal voluntary diving behavior, following a partial to complete recovery from the acid-base disturbance. As previously noted, under conditions commonly encountered in the environment of amphibians and reptiles, oxygen content of water is less than that of air at the same temperature. Water’s O2 carrying capacity is temperature dependent; cold water contains more oxygen than warm water under the same atmospheric pressure. Water and air, however, generally have environmental CO2 concentrations

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that are much lower than those levels commonly found in the blood of an animal occupying that environment. These concentration differences lead to O2 absorption and CO2 release following a typical vertebrate pattern involving lung epithelia. However, CO2 release frequently occurs through supplementary pathways; some CO2 exchange occurs across the surface of the skin. Although reptilian skin is more impermeable because of the scales, significant CO2 exchange can occur across the scale hingeinterscalar spaces. As much as 20 to 30% of CO2 exchange can occur across the skin in this fashion (Zug et al. 2001). Although more predominant in some aquatic reptiles, some terrestrial reptiles (e.g., Lacerta spp.) consistently rely on CO2 loss from the skin (Zug et al. 2001). Respiratory surfaces may represent a significant route of contaminant absorption and uptake, particularly for aquatic organisms. The diversity of respiratory surfaces in reptiles has considerable implications for the absorption of toxic compounds. Depending on the developmental stage or phase of annual activity cycle, reptiles may use different respiratory surfaces, and the differences in membrane surfaces will affect exposure and dose. For example, chemical absortion across skin versus buccopharyngeal mucosa versus cloaca mucosa will affect the extent of chemical uptake from the environment. As noted for observations focused on dietary exposures and their links to gastrointestinal and digestive physiology, and for observations focused on integrative processes related to thermoregulatory, osmoregulatory, and excretory physiology, the range of functional and morphological adaptations associated with gas exchange and respiratory systems in reptiles ensures that general statements regarding exposures in the field must be developed with caution. Beyond the relatively simple appearance of exposure equations (Equations 5.3 and 5.4) and their regulatory applications, there are ample research opportunities for physiological ecologists that would undoubtedly improve our analysis of exposure in herpetofauna.

5.4 Endpoints Commonly Linked to Chemical Exposures to Amphibians and Reptiles in Laboratory and Field Consistent wth our cursory overview of exposure as viewed through the eyes of a physiological ecologist, in the following sections we will briefly focus on effects as outcomes of exposure. Again, the physiological ecologist has long considered outcomes as adaptations linked to selective pressures associated with exposures to environmental stressors such as challenging osmotic environments and extreme temperatures. Similarly, outcomes through the eyes of the ecotoxicologist are commonly called endpoints. While not generally considered within an evolutionary context, outcomes as endpoints share technical foundations with outcomes as adaptations. Indeed, early in its development, ecotoxicology’s relatively close ties to physiological ecology are apparent given the role of chemical stressors such as early-generation chlorinated hydrocarbons as “selective pressures” for development of pesticide resistance in target species and observations of metal resistance in plant species long exposed to soils rich in metallic ores. Research on amphibian ecotoxicology has continued to expand, as indicated by Sparling et al. (Chapter 1, this volume), including an increased number of endpoints that can be measured and interpreted within the context of effects linked to chemical stressor exposure. Although recent additions to the collection of endpoints have been noted (see, e.g., Chapter 14, this volume), common endpoints for evaluating effects in amphibians remain focused on growth and development (including stage-specific and time-specific endpoints related to metamorphosis), behavioral alterations, and biomarkers of exposure (e.g., changes in nucleic acids, enzyme activity, mRNA, protein synthesis) and biomarkers of effects (e.g., limb and skeletal deformities). Similarly, research focused on the ecotoxicology of reptiles has yielded similar lists of outcomes related to exposure and effects.

5.4.1 Growth For both amphibians and reptiles, ecological parameters such as fecundity, juvenile dispersal, adult fitness, and survivability are often dependent on growth and size reached during development and maturation. Although growth (e.g., length as snout-to-vent length [SVL] and body weight) and other

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growth-related endpoints may be evaluated from organismal data collected from the field or from laboratory exposures, baseline values for measurements such as SVL, developmental rate, and size at metamorphosis tend to be more variable under field conditions. Thus, age-specific body-size-related endpoints may not always be sensitive measures to evaluate chemical effects on herpetofauna. Physiological ecologists ply their energetics-based tools (Rice 1990; Widdows and Donkin 1991) to ecotoxicological problems to evaluate growth and reproductive potential in amphibians and reptiles (Rowe et al. 1998). For example, food will be consumed, and energy and materials assimilated, then allocated by priority to maintenance costs supporting basic physiological processes. Energy and materials remaining after maintenance costs are satisfied are then available for growth, reproduction, or storage. Exposure to pollutants can increase energy requirments (e.g., increased metabolic rate or protein synthesis) or can decrease assimilation efficiency (e.g., decrease in foraging efficiency) and create a deficit in the energy balance model (Calow 1989, 1991; Rice 1990; Rowe et al. 1998). The difference in maintenance costs can readily serve as an endpoint. Physiological costs associated with a chemical’s direct effects as a toxicant, or indirect effects as a chemical stressor, can be applied to these bioenergetics models to examine the potential effects that they may ultimately have on growth or reproduction. Depending on the toxicants to which the animal is exposed, mechanisms of toxicity acting at the cellular level will be translated into bioenergetic responses that may then be measured as an integrated effect on the individual. Energy available to offset costs linked to growth, reproduction, and survival may subsequently be derived for populations and communities, if models sufficient to these estimations are available or developed (Rice 1990; Widdows and Donkin 1991). Interpreting growth measurements in amphibians and reptiles is not always straightforward (Petranka 1989; Pfennig et al. 1991). Contaminant research may report both decreased growth (Lefcort et al. 1997) and increased growth depending on the chemical effect and test conditions (Rowe et al. 1998). For example, a chemical may either stimulate growth or stimulate early metamorphosis at a smaller size. Alternatively, a chemical that is lethal to tadpoles in a tank may optimize conditions for density-dependent growth in survivors, or chemicals and test conditions that are stressful to tadpoles may increase body size through edema. Although there is a genetic component to amphibian development through its various life stages — development in ovo, hatchling and development of larvae to juvenile status, then attainment of sexually mature adults — growth is also associated with, and responsive to, environmental factors such as temperature, abiotic conditions of microhabitats, and biological interactions such as predation and competition (Bellis 1962; Jung and Walker 1997; Kiesecker and Blaustein 1998). As in the amphibians, temperature effects on growth and development are similarly expressed in reptiles (see, e.g., Booth 2006). For example, ambient nest temperatures experienced during incubation influence size, shape, color, behavior, locomotor performance, and sex determination in many reptiles. In considering growth as an endpoint in ecotoxicological investigations in the field, one should remember the range of selective pressures that influence the life history of animals, particularly as a function of their unique natural history (Peters 1983), as well as proximate factors that constrain body sizes (e.g., Van Valen 1973; Schmidt-Nielsen 1984; Maurer et al. 1993; Brown et al. 1993). Indeed, in the absence of both laboratory and field studies, if competing factors that influence growth, such as temperature (Huey and Berrigan 1991; Sinervo and Adolph 1994), are not given sufficient consideration in study design, a solitary focus on dietary exposures as dominant factors influencing chemical exposures may lead investigators astray. For example, while measuring growth (e.g., as length in terms of SVL or weight) in the laboratory may be relatively straightforward, in translating those growth endpoints with those same endpoints observed under field settings, investigators must be wary of the wide array of selection forces and mechanisms that influence growth. Body size may be difficult to quantify (Dunham 1978; Gaston and Lawton 1988) relative to exposure conditions related to predation pressure (Owen-Smith 1993) or interspecific competition for food (Illius and Gordon 1987). In the field, body size in reptiles is often limited by food intake, which in turn depends on available forage or prey resources. Differences in supply

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and quality of foods, food intake, and their implications for phenotypic differences in body size may be compared energetically with, for example, field metabolic rates from the literature (Nagy 1982; Nagy and Shoemaker 1984), and with experimental outcomes of feeding trials completed under controlled conditions wherein chemical contaminants are incorporated into diets. By opting to integrate field-laboratory studies, distinguishing between effects of “environmental chemicals” and “other stressors” may be achieved by tracking differences in growth and net energy gains (or losses) between variously challenged animals. From companion field studies, competing factors influencing growth and body size (e.g., predation and interspecific food competition) can be more adequately considered in the interpretation of potential differences in body size based on food availability alone or reduced growth linked to diminished net energy available for growth associated with chemical exposures. Field-oriented physiological ecologists have long appreciated the heterogeneous world in which animals live and have learned to embrace variance. In fact, changes in the variance of physiological responses within populations may be an effect of environmental change. Hochachka and Somero (2002) amply summarized mechanisms for short- and long-term responses to environmental shifts such as temperature, yet studies of responses to changes in food supply (both quantity and quality) remain challenging given the integrative systems (e.g., digestive, hepatic, renal, circulatory, and neurobehavioral) and tissues (e.g., adipose, muscle, and skeletal) that may be involved. Likewise, depending on the level of organization, temperature-related studies focused on these response components may vary as a function of spatiotemporal scales (from local to global, and from seconds to months to years) that strongly influence exposure.

5.4.2 Reproduction and Endocrinology Reproduction is a crucial event in the life history of every organism that ensures continuation of the species. Anything that interferes with reproduction or subsequent embryonic development ultimately may lead to extinction. Because chemical signals (often hormonal or pheromonal) direct reproduction and development, these processes are sensitive to chemical perturbations from the environment. In fully formed adults, disruptions due to environmental chemicals may lead only to activational effects, which are often temporary imbalances. However, for developing embryos, chemical signals direct the formation of anatomical systems and establish physiological set points, and signal disruption due to environmental contaminants during development may lead to permanent organizational level effects, such as anatomical defects or physiological imbalances. Toxicological effects on amphibian and reptilian reproduction can range from prezygotic effects to postzygotic effects. For example, impaired gametogenesis in adults exemplifies a prezygotic effect that reduces fertilization and subsequently the production of hatchlings. Postzygotic effects may contribute to reduced offspring survivorship (e.g., through maternal transfer of contaminants) or through disruption of normal development patterns. Adverse behavioral effects may diminish reproductive success through prezygotic (e.g., by impaired mating displays and reduced abilities to attract mates, or more generally, by impaired timing and type of breeding behavior) or postzygotic (e.g., egg attendance, hiding, guarding, carrying, and feeding offspring [Gross and Shine 1981]) mechanisms. As simply measured by the number of offspring entering the next generation, reproductive success depends on unaffected embryonic development, gender determination, and hatchling growth, all of which commonly vary across a range of species within a genus or family. Within a species, these factors may be highly variable from breeding season to breeding season, depending on environmental conditions potentially affecting reproductive fitness (Highton 1956; Blair 1961; Corn and Livo 1989). Given the variability in various life history traits linked to reproduction, it is best if such information is characterized for the species of concern to the evaluation. This information increasingly is being provided as part of amphibian and reptilian toxicological research (Berrill et al. 1995; Blaustein et al. 1996; Gardner and Oberdörster, 2006), yet remains insufficient to most implementations of the risk assessment process focused on the herpetofauna.

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5.4.2.1 Reproduction and the Environment Reptiles and amphibians have evolved to survive in many biomes, ranging from the tropics to above the Arctic Circle, and from the oceans to the deserts. Within these biomes, they have further adapted for many microhabitats, including terrestrial, fossorial, arboreal, freshwater, and marine. To maximize reproductive output across members of each class, reproductive characteristics vary among species. In most species, reproduction is seasonally timed against extrinsic factors such as temperature, photoperiod, and precipitation in order to optimize survival of the offspring (Karsch et al. 1984; Wingfield and Kenagy 1986). In temperate zones, amphibian breeding is cyclic, gametogenesis is seasonal, and adult gametes mature uniformly. Breeding activity is controlled by endogenous neuroendocrine cycling coupled with extrinsic, seasonally derived triggers such as temperature, photoperiod, and precipitation (Blair 1961). Timing varies geographically and latitudinally, although within a species or population, the cue is constant. In tropical zones and regions of prolonged conditions, such as desert droughts, gametogenesis is continuous; gametes at various stages of development enable some part of the population to be ready to breed at all times (Jacobson 1989). In amphibians, activity may be initiated at any time but typically occurs around rainstorm events. Desert-dwelling Scaphiopus spp. are opportunistic breeders and are physiologically ready to lay their eggs in rain pools at the first major rain event. In temperate reptilian species, reproductive seasonality is typically dependent upon photoperiod and temperature, whereas in tropical species wet and dry cycles may also play a significant role in regulating reproduction (see Palmer et al. 1997 for review). Reproductive cycles range from very short cycles in northern species to nearly continuous reproduction in some tropical species (Fitch 1970, 1982; Moll 1979; Duvall et al. 1982). Species with wide geographic ranges are subject to local environmental variation in conditions. Consequently, these species exhibit within-species variation in reproductive patterns. For instance, some sea turtles (Chlonia mydas, C. depressa) nest throughout the year in tropical portions of their range, but seasonally in temperate regions (Grace 1997). Our understanding of the reproductive endocrinology of reptiles and amphibians continues to advance, yet does not match that of endotherms. Only a few species have been intensively studied. For the vast majority of reptilian and amphibian species, little or no information is available regarding reproductive endocrinology or physiology (Palmer 2000). 5.4.2.2 Pineal Gland The pineal gland (epiphysial gland) is the principal organ for detecting environmental cues and translating them into endocrine signals for regulating reproductive cycles. The pineal gland is absent in crocodiles (Roth et al. 1980), yet occurs in most fishes, amphibians, and reptiles as a saclike diverticulum of the third ventricle of the brain. In anamniotes and lizards, the basal portion of the pineal gland is photosensory, while in other reptiles the gland connects to the suprachiasmatic nucleus from which it receives photoperiod information from the eyes (Quay 1979; Collin and Oksche 1981). During periods of darkness, the pineal gland utilizes N-acetyltransferase and hydroxyindole-O-methyltransferase to synthesize serotonin and then melatonin from tryptophan (Quay 1974). Melatonin also may be produced by the retina in amphibians and reptiles (Ralph 1980; Pang and Allen 1986), and concentrations of plasma melatonin vary diurnally, having greater concentrations during the dark phase (scotophase) than during the light phase (photophase). The cyclical level of melatonin has several functions, including regulating circadian rhythms, influencing body coloration by changing melanophore size, and regulating annual reproductive cycles by influencing gonadotropin release (Underwood 1992). By linking reproductive events with environmental factors, offspring are more likely to emerge during favorable conditions (Marion 1982). 5.4.2.3 Hypothalamus and Pituitary The pineal gland responds to the external environment, functioning as a transducer of the physical environment linked to endocrine signals (melatonin levels). Melatonin, in turn, influences the

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hypothalamus and subsequently the pituitary, which together regulate the reproductive organs, the ovaries or testes. The hypothalamus secretes gonadotropin-releasing hormone, which in turn stimulates the adenohypophysis of the pituitary to produce the gonadotropins luteinizing hormone (LH) and follicle-stimulating hormone (FSH). Although well characterized in endotherms such as mammals, the functions of endogenous LH and FSH in reptiles and amphibians remain obscure. In turtles and crocodilians, both FSH-like and LH-like molecules have been detected (Licht and Papkoff 1974; Ishii 1991), but squamates appear to rely on a single FSH-like gonadotropin, which may have both LH and FSH activities (Licht 1979). In hypophysectomized Anolis spp., exogenous FSH maintains testicular weight, whereas both exogenous FSH and LH stimulate increased androgen levels (Licht and Pearson 1969). Until conspecific gonadotropins are regularly available for the study of reproduction in reptiles, hormonal function and its role in regulation will have to rely on administration of gonadotropins from other vertebrates.

5.4.3 Female Reproduction 5.4.3.1 Vitellogenin Amphibians and reptiles produce macrolecithal eggs, in which a large yolk serves as an energy reserve for the developing embryo. In nonmammalian vertebrates, yolk production is initiated in the liver of adult females with the production of vitellogenin (Ho 1987). Once synthesized and released by the liver, vitellogenin enters systemic circulation, and is subsequently taken up by developing oocytes and converted to egg yolk (lipovitellins and phosvitins). Structurally, vitelloginin is a phospholipoglycoprotein precursor of egg yolk, and across a range of species its molecular weight commonly ranges from 200 to 250 kDa. In the systemic circulation, vitellogenin usually occurs as a dimer of 400 to 500 kDa (Callard and Ho 1987). Multiple forms of vitellogenin are present in the plasma of several species, including Xenopus, chicken, and several fish species, but each is glycosylated and contains about 1.4% carbohydrate by weight (Ho 1987; Lazier and MacKay 1993). Estrogen is the primary stimulus for vitellogenin production. Estrogen promotes vitellogenesis in members of all nonmammalian vertebrate classes (Ho 1987). In each of these groups, the time of vitellogenin production in adult females corresponds to the period of elevated estrogen levels. Vitellogenesis can be induced in males and in nonvitellogenic females by administration of estrogen. The production of vitellogenin in response to estrogenic compounds is rapid, sensitive, and dose dependent (Ho et al. 1981). Vitellogenin can be utilized as a biomarker for exposure to xenobiotic estrogens (Palmer and Palmer 1995; Palmer et al. 1998; for review, see Palmer and Selcer 1996). Estrogen or estrogen mimics are the sole inducers of vitellogenin (Tata and Smith 1979), although other factors have a role in modulating the vitellogenic response (Ho et al. 1981; Wangh 1982). The strength of the vitellogenic response indicates the relative ability of xenobiotic compounds to stimulate estrogenic pathways. Additionally, through vitellogenesis, transfer of lipophilic contaminants from female fat stores or her proximate diet can expose developing embryos to potentially hazardous concentrations, leading to developmental and other effects (see Rowe 2008). 5.4.3.2 Ovarian Structure and Function The amphibian ovary is a hollow, sac-like structure. The ovaries of reptiles are saccular or membranous structures in which enlarged follicles are prominent. Among crocodilians and chelonians, the membranous ovaries are symmetrically positioned ventral to the kidneys. Among lizards, the ovaries may be either symmetrically or asymmetrically positioned, with one ovary more anterior than the other. In snakes, asymmetrically placed ovaries are the rule, with the right ovary usually larger and more anterior than the left, which corresponds to the length of the adjacent oviduct. The left ovaries of snakes may be reduced or undeveloped. The female ovary consists of oocytes surrounded by both granulosal and thecal layers, with the granulosa layer serving as the primary site of estrogen synthesis (Callard and Ho 1980; Callard and

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Kleis 1987). Estrogen stimulates development of the oviduct in reptiles and amphibians, although outcomes of that development vary across species. In squamates, for example, pyriform cells have been identified in the granulosa, which may be involved in early oocyte development (Uribe et al. 1995). Once vitellogenesis begins, the pyriform cells degenerate, and the theca surrounds the granulosa layer. During development, the thecal and granulosal layers become separated by the acellular membrana propria (Uribe et al. 1995); the theca interna remains glandular, whereas the theca externa becomes fibrous. Rising levels of gonadotropins stimulate follicular recruitment (Palmer et al. 1997), and in reptiles, interstitial glands may form from atretic follicles and exhibit an endocrine function. While ovulation in amphibia is under the control of LH and progesterone, the endogenous stimulus for ovulation in reptiles is unknown. However, administration of exogenous FSH can induce ovulation in Anolis (Jones et al. 1988). In oviparous amphibians, ova are stored until released. Ova released during ovulation subsequently enter the oviducts where egg jellies are deposited. Following ovulation, the granulosal and thecal cell layers of the follicle are transformed into corpora lutea, which secrete progesterone. In amphibians, progesterone induces responsiveness of the oviducts to arginine vasotocin (AVT), with AVT acting to stimulate oviductal contractions during oviposition. Once eggs are fertilized and deposited, the perivitelline chamber increases in volume, as a result of the uptake of surrounding water and accumulation of waste products. The outer membrane is water permeable, but is sensitive to pH outside the neutral range (Dunson and Connell 1982). In reptiles, corpora lutea have been shown to exhibit 3β-hydroxysteroid dehydrogenase activity and synthesize progesterone (Klicka and Mahmoud 1972, 1973, 1977; Licht and Crews 1976). Following ovulation, egg yolks enter the paired oviducts where fertilization occurs. Fertilization is internal in all reptiles and presumably occurs in the anterior oviducts prior to deposition of any egg coats (Palmer and Guillette 1988). In oviparous species, albumen and eggshell layers are deposited within the oviducts. Albumen is a complex mixture of water-soluble proteins that may influence embryonic development (Palmer and Guillette 1991), and is deposited by the anterior glandular portion of the oviduct (Palmer and Guillette 1988, 1991). Subsequently, the eggshell is deposited by the uterus (Guillette et al. 1989; Palmer and Guillette 1990; Palmer et al. 1993), and consists of an inner fibrous portion and an outer calcareous layer (for review, see Schleich and Kästle 1988; Packard and DeMarco 1991). In turtles and squamates, the uterus is homogeneous and produces both the fibrous and calcareous eggshell layers from endometrial glands along its entire length (Palmer and Guillette 1988; Guillette et al. 1989; Palmer et al. 1993). In crocodilians, the uterus is divided into morphologically distinct fiber-producing anterior and calcium-secreting posterior regions (Palmer and Guillette 1992). Oviductal function is regulated primarily by estrogen and progesterone, although androgens such as testosterone and dihydrotestosterone (DHT) also may play minor roles (for review, see Palmer et al. 1997; Selcer and Clemens 1998; Selcer et al., 2005). The oviducts also have been shown to possess specific androgen receptors and aromatase (Smith et al. 1995). Progesterone in reptiles inhibits muscular contractions of the uterine walls by blocking formation of arginine vasotocin (AVT) receptors, preventing early parturition or oviposition (Guillette et al. 1991a,b; Guillette et al. 1992). Corpora lutea typically persist until oviposition or parturition. Declining progesterone levels associated with the involution of the corpora lutea release the inhibition on parturition and oviposition (Guillette et al. 1991b). AVT induces smooth muscle contraction by stimulating local synthesis and release of prostaglandins (PGs) from the uterine wall. The combination of AVT and PGs leads to peristaltic waves of muscular contraction that expels the eggs or embryos (Guillette et al. 1990, 1991a, 1991b, 1992). 5.4.3.3 Reproductive Strategies Amphibians and reptiles show extremely diverse reproductive strategies, including oviparity, viviparity, ovoviviparity, and various degrees of parental care (Guillette 1987, 1989, 1991; Hanken 1989; Wake 1993). Caecilians, salamanders, and frogs display both oviparity and viviparity. All turtles and crocodilians are oviparous, but viviparity has evolved numerous times

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among the squamata (Blackburn 1982, 1985; Shine 1985; Shine and Guillette 1988). In viviparous reptilian species, the embryo implants into the uterine lining, and the oviductal glands that secrete the eggshell membranes degenerate prior to implantation. The thin uterine lining becomes highly vascularized and strongly co-mingled with embryonic tissues to form a placenta (Weekes 1935; Yaron 1985; Stewart and Blackburn 1988; Stewart 1992). Most nutrients are supplied to the embryo in the form of yolk provided by the ovary. However, significant gas exchange occurs across the placenta. Parthenogenesis has been reported in some lizard and snake species. Examples are found among species of Sceloporus, Cnemidophorus, Ramphotyphlops, Elaphe, Agkistrodon, Gymnophthalmus, and some geckos (Burgin et al. 1997). Their ova initiate development without the presence of males, and the unfertilized ova develop normally into exact copies of the mother. Some salamanders of the genus Ambystoma are gynogenetic. Gynogenesis is similar to parthenogenesis, with the exception that the eggs must be stimulated by the presence of sperm in order to start development. Both parthenogenesis and gynogenesis reduce genetic variation within the population and thus limit the population’s ability to adapt to new environmental challenges, such as the introduction of emerging contaminants. Annual fecundity in amphibians can range from 1 to more than 80,000 viable hatchlings per clutch (Jørgensen 1992; Baker 1992). Larger species and those with generalized modes of egg production deposit larger clutches. Generally, less than 5% of eggs will survive to metamorphosis. Among reptiles, the fecundity can range from 1 to over 100 eggs per clutch for sea turtles.

5.4.4 Male Reproduction 5.4.4.1 Testis Structure and Function The testis of urodeles is of the cystic type, similar in structure and function to those of fishes. The uredele testis consists of 1 or more lobes, each containing several ampullae, which in turn are composed of several germinal cysts. Germ cells within a cyst divide synchronously, so that all sperm mature within a cyst at the same time. The testis may have cysts in various stages of development, representing differing reproductive episodes, such as temporally separated breeding events. The urodele testis exhibits Sertoli cells and lobule boundary cells, which both express steroidogenic activity. The anuran testis is structurally more similar to that of amniotes, consisting of seminiferous tubules with a permanent germinal epithelium and conspicuous interstitial tissues. The interstitial cells are steroidogenic and produce androgens. The seminiferous tubules contain Sertoli cells, which regulate sperm production and are also steroidogenic. In male reptiles, the seminiferous tubules of the testis are the functional units of reproduction, and testicular recrudescence is stimulated by rising levels of gonadotropins (Licht et al. 1977; Licht 1979; Ishii 1991). Sertoli cells are present within the seminiferous tubules among the developing sperm cells, whereas interstitial Leydig cells are found between the tubules. Both Sertoli and Leydig cells secrete androgens (Mahmoud et al. 1985). Following completion of spermatogenesis, sperm migrate into the epididymis where they are stored until subsequent release. In contrast to other orders in the class, male squamates have an accessory reproductive organ known as the renal sex segment, which is part of the kidneys, developing from the uriniferous tubules (Prasad and Reddy 1972). Under the influence of androgens, the tubules of the renal sex segment hypertrophy and become engorged with secretory granules. These secretions become mixed with the semen during ejaculation and may function in maintaining sperm viability, although their specific function remains unknown. Some authors have suggested that the sex segment is homologous to the seminal vesicles of mammals (Norris 1997). 5.4.4.2 Fertilization and Copulatory Organs Fertilization may be either internal or external, depending upon species. Anuran mating generally occurs by way of cloacal apposition, wherein eggs are laid in water, and fertilization is external.

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Though there are salamanders that exhibit external fertilization, 90% of the species fertilize eggs internally. Some terrestrial salamanders breed in the fall and store sperm internally in spermatheca or dorsal diverticulum of the female cloaca and in urogenital pouches (Massey 1990), and eggs are deposited and fertilized the following spring. Ambystomid males deposit spermatophores (cloacal gland secretions of sperm packets encased in a jelly-like substance) in ponds (Bishop 1941), where females subsequently pick up the spermatophores and deposit eggs. Caecilian species reproduce biennially, and exhibit intromission and internal fertilization (Jørgensen 1992; Wake 1993). Caecilians possess an elaborate intromittent organ, the phallodeum, associated with the posterior portion of the cloaca. Fertilization in reptiles is always internal, due to production of shelled amniotic eggs or live young. Male reptiles possess copulatory organs used to transfer sperm to the female. In turtles and crocodilians, the copulatory organ is a single penis, and in squamates, paired hemipenes are present. These organs are housed within the cloaca until intercourse, at which time they become engorged with blood and extend through the cloacal opening.

5.5 Reproductive Ecology 5.5.1 Parental Care Parental care is rare among caecilians, and is found in some anurans (Townsend et al. 1991) and in many, if not the majority of, urodeles (Tilley 1972). Parental care includes attending eggs, transporting eggs or young, and feeding young. Depending on species, duration of parental care can extend to more than 150 days (Duellman and Trueb 1994) and can involve either the male (e.g., in Hynobiidae) or female (e.g., in Plethodontidae, Desmognathus fuscus and Aneides lugubris). For review and species references, the reader is referred to Duellman and Trueb (1994). Parental care is absent among turtles and tortoises, which all bury their eggs, but is found in some lizards, snakes, and the crocodilians. Some skinks guard their eggs and presumably aid in maintaining adequate moisture levels during incubation. Several pythons not only guard their eggs, but actively increase incubation temperature by active shivering thermogenesis, thereby shortening incubation time. Crocodilians exhibit extensive care of both offspring and eggs by guarding them aggressively from predators and assisting the young to emerge for the nest mound. Studies indicate that, in many cases, parental care improves rate of hatching and survival. The energetic cost of care depends on the extent to which parents continue feeding and the frequency of egg deposition. Underlying physiological and endocrine mechanisms associated with these diverse strategies need additional study (Townsend et al. 1991), and contaminant effects, direct with respect to care or indirect with respect to the energetic costs of care, have yet to be investigated.

5.5.2 Offspring Survival To a large extent, growth and posthatch activities of offspring are environmentally determined. For example, many studies indicate that large-bodied offspring exhibit increased survival and are more fit than their small-bodied cohorts (John-Alder and Morin 1990; Platz and Lathrop 1993); hence, some anurans (e.g., Rana catesbeiana and R. clamitans) overwinter as tadpoles to metamorphose in the spring when their body mass would be greater. If not overwintering, these species would be more susceptible to predation as tadpoles, and the risk of their ponds evaporating prior to their metamorphosis would be increased (Anderson et al. 1971; Cooke 1973; Bishop 1992; Hota 1994; Bridges 1997). In amphibians, overall ecological and physiological costs and benefits of remaining aquatic or metamorphosing to a terrestrial adult depend on the relative quality of the pond and surrounding terrestrial habitat. Further, studies using allometric engineering in reptiles indicate that alterations in the quantity of yolk may alter offspring size (Sinervo et al. 1992). Production of vitellogenin is also susceptible to environmental endocrine disruptors (Palmer and Palmer 1995; Palmer et al. 1998), which may alter egg or clutch size (Irwin et al. 2001). Ecotoxicological implications

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of altered offspring size or delayed or reduced growth resulting from contaminant exposure need to be considered at the population level and within the field environment (Smith 1987; Figiel and Semlitsch 1990; Pfennig et al. 1991; Whiteman et al. 1996).

5.5.3 Lifespan and Exposure Very little information is known about basic life history and reproductive potential of most amphibians and reptiles — even for the more commonly studied groups (Matson 1998). The lifespan for amphibians ranges from less than 1 year to more than 30 years (Bufo americanus; Carey and Judge 2000), and that of reptiles to well over 100 years (Cary and Judge 2000). Information on reproductive potential and duration (e.g., age at sexual maturation, average lifespan) is vital to evaluating contaminant-induced die-offs and other catastrophic events on population recovery over time. For those organisms that deposit many thousands of eggs, information on recruitment and reproductive strategies may be far more important than the clutch size or SVL when evaluating ecotoxicological effects (Larson 1998).

5.6 Development Within field settings, development of amphibians and reptiles from early embryo to adult occurs as a series of events sensitive to environmental cues. As key factors contributing to the normal developmental process, these environmental cues have acted through evolutionary time and have influenced development to yield the wide range of phenotypes manifested as varying expressions of the genotypes characteristic of the herpetofauna. The environment can affect development in several ways, ranging from cued events normally experienced by organisms during their ontogeny to those interactions with environmental stressors that are newly encountered or encountered under conditions that were not previously experienced in the animal’s phylogenetic history. For example, through evolutionary time seasonal cues such as photoperiod, temperature, or hydration may alter an organism’s development to increase its fitness, yet newly encountered environmental stressors may contribute to disruption of the normal developmental process for a species. As such, both physiological ecologists and ecotoxicologists view developmental events characteristic of the herpetofauna as outcomes potentially linked to exposures to environmental stressors encountered in the field. However, there may be time frame differences between the natural stressors and anthropogenic stressors, which typically appear much more quickly in the environment. Embryonic development is one of the more sensitive stages in the life of amphibians and reptiles. Yet, the role of exposures to environmental chemicals is not completely understood and continues to be the subject of study. Since the anatomical and physiological systems developing in embryonic reptiles are controlled by numerous, and oftentimes interacting, chemical signals, adverse effects potentially linked to exogenous chemicals should be more sufficiently characterized without undo reliance on comparative analyses focused on endotherms or a few species of amphibians or reptiles that are regarded as representative of their class. Chemicals in the environment can alter either the signals themselves or the animals’ ability to recognize them.

5.6.1 Sex Determination Sex determination is linked to several different developmental mechanisms. Genotypic sex determination (GSD) results from the genetic makeup of the embryo, frequently manifested by differences in sex chromosomes. The homogametic sex will have 2 identical sex chromosomes, and the heterogametic sex will have 2 different sex chromosomes. In mammals and many anuran amphibians, the female is the homogametic sex designated by 2 X chromosomes (XX), and the male is the heterogametic sex designated by an X and a Y chromosome (XY). Birds and many urodele

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amphibians present males that are homogametic (designated ZZ), while females are heterogametic (designated ZW; see Norris 1997). Temperature-dependent sex determination (TSD) is encountered in all crocodilians, many turtles, some lizards (geckos and lacertids), and perhaps some amphibians (Norris 1997; Zug et al. 2001). TSD is characterized by a relatively narrow range of temperatures that affect sex determination. Temperature sensitivity usually occurs in a relatively limited time frame, generally in the first to middle third of the incubation period (Vogt and Bull 1982; Bull et al. 1990; Haig 1991; Desvages et al. 1993; Spotila et al. 1994). There is considerable variation among reptiles regarding the temperature that determines a particular sex, and the critical temperature at which both sexes are produced. In most species, there is an “all or none” effect in which only males or only females are produced on either side of the critical temperature range. There are 3 general patterns of TSD in reptiles (Bull 1980). In some turtles, males are produced at low temperatures (generally less than 25 to 28 °C), and females are produced at high incubation temperatures (usually greater than 31 to 33 °C). Intermediate temperatures produce a gradation of males and females (Vogt and Bull 1982; Ewert and Nelson 1991). In contrast, in some lizards lower temperatures produce females and higher temperatures produce males (Bull 1987). Different yet are the crocodilians, some turtles (Chelydra spp.), and some geckos (e.g., Eublepharis macularius and Hemitheconyx caudicinctus) in which males are produced at intermediate temperatures and females at both high and low incubation temperatures (Webb and Cooper-Preston 1989; Ewert and Nelson 1991; Viets et al. 1993; Lang and Andrews 1994). In the wild, significant temperature variation can occur among nesting localities, resulting in sex ratio variation of recruits (Vogt and Flores-Villela 1992). Over the long term and considering large numbers, these deviations tend to balance out and produce roughly equal numbers of males and females (see reviews in Bull 1980; Janzen and Paukstis 1991; Lang and Andrews 1994; Viets et al. 1994). Sex steroids and the metabolism of sex steroids play a role in TSD. Administration of estrogen or exogenous estrogens at male-inducing temperatures in reptiles can reverse the production of males and lead to a higher percentage of females than normal or even completely sex reverse the embryos, producing all females (Raynaud and Pieau 1985; Wibbels et al. 1994). Compounds that interfere with estrogen synthesis or action may disrupt ovarian development or even induce testis formation at female-producing temperatures (Lance and Bogart 1991, 1992; Wibbels and Crews 1992, 1994; Crews et al. 1994; Pieau et al. 1994a; Richard-Mercier et al. 1995). This indicates that estrogens play a major role in sexual differentiation in reptiles with TSD. Temperature also influences the synthesis of estrogen by influencing the activity of aromatase, the enzyme responsible for the formation of estrogen (Crews 1994; Jeyasuria et al. 1994; Jeyasuria and Place 1997). In turtles (Desvages and Pieau 1992; Pieau et al. 1994b) and alligators (Smith 1997), aromatase is active only at femaleproducing temperatures, which suggests that aromatase activity and the production of estrogen may stimulate female development in reptiles with TSD. Temperature effects on androgen synthesis and activity may also contribute to TSD in reptiles. For example, administration of testosterone has little effect in inducing males in species with TSD, but DHT can induce predominantly male hatchlings at temperatures that would normally produce both sexes in Trachemys scripta (Crews et al. 1994). DHT is produced from testosterone via the action of 5-alpha-reductase, and inhibitors of 5-alpha-reductase demonstrated that the enzyme plays a role in testis differentiation and ultimately in production of males (Crews and Bergeron 1994). Although additional study must be completed, species-specific differences in the synthesis and activities of estrogens and androgens in species with TSD may determine an embryo’s sexual differentiation. The ability of environmental factors to alter reptilian sex ratios may be critical to population level responses to chemical stressors, and altered outcomes linked to TSD-chemical stressor interactions may have significant implications for ecotoxicology (Crain and Guillette 1998). Despite increased knowledge gained since Palmer’s original overview (Palmer 2000), questions regarding the mechanisms of TSD in reptiles remain an active area of physiological research. The role that

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chemical stressors — singly or in concert with other environmental stressors — play in exposures to reptiles in the field requires additional study. In anurans, initial sexual differentiation and development of secondary sex characteristics during maturation are at some level dependent on aromatase activity and exposure to steroids (estrogen or androgen) and thyroid hormones (see Hayes 1997 for review; Ertl and Winston 1998). In amphibians, metabolism generally is temperature dependent (Hayes and Licht 1995), and since aromatase activity is influenced by temperature, alterations in estrogen and testosterone levels may also be induced by temperature. Gonadal reversal experiments conducted with Pleurodeles, similar to those conducted with reptiles (Bergeron et al. 1994; Crews et al. 1996), successfully demonstrate a thermosensitive window; however, test conditions were not considered environmentally relevant (Pieau et al. 1994b; Chardard et al. 1995). Although effects on sex characteristics observed under field conditions could be related to endoctrine disrupting chemicals (EDC) exposure that results in altered endogenous androgenic or estrogenic receptor binding or function, observed effects could also result from the activity of the temperature-sensitive steroidal enzyme aromatase (Pieau et al. 1994b; Sheffield et al. 1998; Ertl and Winston 1998). Hormonal and reproductive measures under baseline or controlled conditions are potential endpoints for evaluating EDC effects in animals tested during laboratory and microcosm studies, and in those collected from contaminated field conditions (Gendron et al. 1997). Hormone disruptors, such as estrogenic or androgenic EDCs, can affect critical life stages, for example, the organization of gender determination of the gonads and brain during initial development and the activation of endocrine and behavioral responses during sexual maturation (Noriega et al. 1997). For example, brain neurons associated with the male frog larynx are sexually dimorphic; treating females with androgens can masculinize their larynx (Burggren and Just 1992). Some EDCs, such as dioxins, also target the thyroid system, and may therefore have effects on sexual differentiation and other developmental processes, including metamorphosis. Research on the role of environmental factors in sex determination and on the activation and timing of the interaction between steroid and thyroid hormones continues.

5.6.2 Metamorphosis Most amphibians follow a developmental process unique to vertebrates — metamorphosis. Metamorphosis consists of a series of postembryonic biochemical, morphological, and physiological changes that transform larvae into adults (Dent 1988; Galton 1988; Eales 1990; Hayes and Licht 1995; Kaltenbach 1996; Wright et al. 1997; Denver 1998). Not all amphibians metamorphose, however. Most Plethodontidae salamanders and some anurans deposit eggs on land where embryos undergo direct development, effectively bypassing the larval stage. Neotenic mudpuppies (Necturus spp.) and hellbenders (Cryptobranchus spp.) lay their eggs in water and remain aquatic throughout their life. When amphibians do metamorphose, the changes that occur are significant (Frieden 1963; Kaltenbach 1996). Stages during metamorphosis are defined by hormonal and anatomical events such as tail resorption or skin keratinization. The thyroid plays a critical role in regulating metamorphosis. During premetamorphosis, follicular cells of the paired thyroid glands grow and become secretory (Gancedo et al. 1997). Tetraiodothyronine (T4, thyroxine) and triiodothyronine (T3) are released into the bloodstream, stimulating an increase in the peripheral thyroid receptors (Wright et al. 1997). Subsequent to this stage, T4 and T3 levels increase and the hypothalamic-pituitary-thyroid axis is activated (Norris and Gern 1976). Thyroid-stimulating hormone (TSH) from the anterior pituitary acts on the thyroid to control gland activity (Norman and Norris 1987; Miranda et al. 1996). During metamorphic climax, thyroid hormones (primarily T3) induce the biochemical, morphological, and functional changes associated with the transition to adulthood (Frieden 1963; Dent 1988; Galton 1988). Direct and indirect endocrine functions include the thickening and keratinizing of the thin, multilayer larval skin, which helps to conserve water and reduce the potential for serious mechanical injury in

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the terrestrial adult. Mucus produced by dermal glands helps in thermoregulation and osmoregulation, forming a barrier to epithelial water loss (Shoemaker et al. 1992; Duellman and Trueb 1994). Thyroid hormones influence ossification of cartilage and myosin/tropomyosin synthesis for muscle (review in McNabb and King 1993). Thyroid activates lipogenesis, resulting in lipid storage in the liver and fat bodies to meet increased energy needs. During metamorphosis, the tail is resorbed, gill arches degenerate, and gills regress (Duellman and Trueb 1994). Following tail loss, thyroid hormone levels decrease (Boutilier et al. 1992). Other hormones involved in metamorphosis include prolactin, growth hormone, insulin, and adrenal corticoids (Brown et al. 1991; Kobayashi and Kikuyama 1991; Hayes and Wu 1995; Hayes 1995a, 1995b; Kloas et al. 1997). Anterior pituitary prolactin stimulates tissue growth (e.g., tail and gut in tadpoles and gills in salamander), regulates water and electrolytes, stimulates intestinal absorption of amino acids and glucose, and decreases hydrolytic enzyme activity involved with tissue regression (Dent 1988). Lipolytic prolactin levels run counter to thyroid hormone concentrations, possibly regulating at the level of the brain nerve terminals and monoaminergic system rather than at the receptor level (Burggren and Just 1992). Adrenal steroids accelerate thyroid-induced metamorphosis, whereas growth hormone promotes tissue growth. Pancreatic insulin stimulates cutaneous ion transport (Boutilier et al. 1992) and helps control blood glucose levels. Insulin levels increase in the pancreas and serum in the larvae until metamorphic climax. Each of the transition states of metamorphosis — egg to larva to adult — presents different interactions with the environment; hence, exposure across life history stages potentially varies significantly. The disposition of chemicals accumulated in an earlier stage of development may affect stages yet to come; for example, chemicals stored in the larval tail may be redistributed and become available for metabolism, and epithelial restructuring can modify rate and transport during chemical uptake. Historically, both laboratory and field studies have indicated varying sensitivities to chemical exposure across stages of development (Sanders 1970; Saber and Dunson 1978; Dial and Bauer 1984; Dial and Dial 1987; Berrill et al. 1993). Factors contributing to differential sensitivity include individual tolerance (Dial 1976; Dial and Bauer 1984; Dial and Dial 1987; Rowe et al. 1998), development of resistance (Browne and Dumont 1979), time of exposure relative to organogenesis and metabolic state of the embryo (Honrubia et al. 1993), differential development of immune or other physiological resistance response mechanisms (Dial and Dial 1987; Sheffield et al. 1998; Van Der Kraak et al. 1998), and the ability to modify chemical uptake, metabolism, or clearance due to temperature regulation, degree of hydration, or protein and enzyme synthesis and induction (Suzuki and Akitomi 1983; Rosenbaum et al. 1988; Herkovits and Oerez-Coll 1993; Lizana and Pedraza 1998). The extent and permanence of adverse effects depends on the timing of exposure during cellular development (Honrubia et al. 1993) or on the stage-dependent ability to synthesize effective isoforms of proteins such as metallothionein (Herkovits and Oerez-Coll 1993; Vogiatzis and Loumbourdis 1998). Many of the more persistent lipophilic chemicals may be sequestered in lipid-storing premetamorphs, but they may be redistributed in the carbohydrate-storing postmetamorphs (Honrubia et al. 1993). The egg stage cannot completely avoid chemicals that may occur in their aquatic environment; however, depending on the chemical, envelope and jelly coatings may confer some protective barrier to the developing embryo (Berrill et al. 1997; Jung and Walker 1997). The extent of this protection may depend on the number and type of envelopes and on the distribution of eggs and egg mass design (Dunson and Connell 1982; Seymour and Bradford 1995; Carey and Bryant 1995; Ovaska 1997). Risks may differ depending on exposure; for example, aquatic amphibians cannot completely avoid dissolved chemicals, whereas terrestrial ones may be able to avoid contaminated microhabitats behaviorally (e.g., by burrowing). For example, postmetamorphic juvenile Scaphiopus couchii are more susceptible to herbicide toxicity than are their adult counterparts, possibly because their smaller size and larger surface-to-volume ratio lead to increased chemical uptake (Judd 1977). On the other hand, late-stage larvae of both Rana pipiens and Bufo americanus are more sensitive to herbicides than are those tested at an earlier stage (Howe et al. 1998). Interspecific sensitivity may

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be related to size of eggs and developmental stage (Berrill et al. 1997). Because of differences in limb development, chemicals such as tributyltin are effective teratogens on hindlimb development but not on limb regeneration (e.g., in Ambystoma mexicanum [Chang et al. 1976; Scadding 1990]). EDCs can alter the timing and rate of metamorphosis, the development of immune and stress responses, and the overall fitness of the transformed population (Bishop 1992; Bonin et al. 1997; Larson 1998; Rohr et al. 2003, 2004, 2006). In addition, exposure of larvae to EDCs can have insidious and persistent effects on adults (Rohr and Palmer 2005). In some instances, species may become reproductively mature without completing metamorphosis. Pedomorphic larvae are sexually mature as a result of accelerated gonadal development (Pough 1989; Whiteman 1994; Whiteman et al. 1996). Neotenic larvae are able to reproduce as a result of delayed somatic development. Under normal conditions, neoteny does not occur in anurans, and obligate neoteny in species of Necturus, Proteus, and Amphiuma is related to tissue insensitivity to thyroid hormones (Norris et al. 1977; Hayes 1997; Larson 1998). Facultative neotenic populations such as Ambystoma tigrinum occur under certain environmental conditions. One neotenic morph remains aquatic throughout its life cycle and lives in a permanent pond environment; a smaller aquatic morph inhabits more ephemeral ponds and metamorphoses to a reproducing adult; and a third morph remains aquatic but morphologically develops a larger head, wider mouth, and longer teeth to become a more significant carnivorous predator (Collins 1981; Whiteman and Howard 1997).

5.6.3 Endocrine-Disrupting Compounds As demonstrated previously, reproduction and development are strongly regulated by the endocrine system and susceptible to the effects of EDCs (see Hayes 2000; Guillette 2000). Recent studies with wildlife indicate that other endocrine systems can be critically impaired by exposure to EDCs, independent of direct receptor-binding interference and outside of initial development and metamorphosis. Additional systems and hormones potentially affected in amphibians and reptiles include, but are not limited to, the following: • neuroreceptors associated with the pineal gland, pheromones, and sensory organs used for intraspecific communication; • nonsteroid hormones such as GnRH and arginine vasotocin related to reproduction and mating behavior; • pituitary melanophore-stimulating hormone and its control of skin pigmentation; • parathyroid hormones, prolactin, and vitamin D related to Ca regulation; • pancreatic insulin and glucagon regulation of glucose and fat deposition associated with seasonal activities; and • catecholamines that stimulate glycogenolysis and control the cardiovascular system and adenocorticotrophin and glucocorticoids (e.g., corticosterone) associated with quick response and long-term adjustments to stress or environmental change (see discussion in Honrubia et al. 1993; Hayes et al. 1997). Any disruption to normal endocrine activity needs to be understood within the context of life stage and potential intrinsic interactions. Reference hormonal levels and their influence may be different in larval and adult animals (Duvall and Norris 1980; Kwon et al. 1991, 1993; Hayes and Licht 1995; Hayes and Wu 1995; Hayes 1995a, 1995b; Hopkins et al. 1997; Kloas et al. 1997) and under different environmental conditions (Johnson 1976). Differences may also be expressed among species. For example, T4 is involved in molting in salamanders but not in anurans (Herman 1992), and as a result, sensitivity and response to EDCs will likely differ. The endocrine interactions among species, developmental stage, reproductive condition, and the environment need further research in order to accurately interpret both laboratory and field experimental data.

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5.7 Behavior Changes in behavior are often the first indication of exposure to environmental stressors, including chemical contaminants. For example, a common observation used in laboratory studies with amphibians is avoidance response to gentle prodding; for example, unexposed tadpoles will move directly away from prodding (Rosenbaum et al. 1988; Walker et al. 1996; Sparling et al. 1997). Observed swimming patterns can be indicative of central-neural, peripheral-neural, or neuromuscular system effects. Observations can be quantified; swimming speed for individuals in Bufo americanus and Rana clamitans has been a useful measure of contaminant effects (Jung and Walker 1997; Raimondo et al. 1998). Although there are some recent data on indirect toxic effects on predation (Jung and Walker 1997; Raimondo et al. 1998; Relyea and Mills 2001; Relyea 2003, 2004, 2005; Rohr and Crumrine 2005), to date very little data have been recorded on effects of contaminants on activities such as amplexus, calling, brooding and parental behavior, ability to catch prey, or level of seasonal migratory drive. Similarly, although the behavior of reptiles, particularly those activities linked to integrative responses associated with hormonal control of reproduction, continues to be a research area of keen interest to herpetologists and evolutionary biologists, very little work has focused on the role that environmental chemicals may have in modulating behavior with the exception of work related to EDCs in the environment.

5.7.1 Sensory Organs Linking external environmental stimuli to responses observed in herpetofauna in field or laboratory settings hinges, in part, on sensory organs characteristic of amphibians and reptiles (Gans and Crews 1992). As Kardong (2005) suggests, the sensory systems of amphibians and reptiles typically present structures and functions similar to other vertebrates, although species-specific differences across the range of animals contribute to much variation on the basic vertebrate motif. Sense receptors may be simply categorized as somatic receptors (e.g., neuromast organs, the membranous labyrinth of the inner ear, light receptors, proprioreceptors, and capsulated and uncapsulated cutaneous receptors) and visceral receptors. Modifications of these basic types occur in some of the herpetofauna, such as infrared receptors of snakes. As in vertebrates across all classes, visceral receptors include olfactory organs, taste buds, and the vomeronasal organ, as well as naked nerve endings in the viscera that serve as stretch receptors, chemoreceptors, baroreceptors, and osmoreceptors. Sensory organs may be characterized as being relatively widespread, serving a general function such as sensation of temperature or proprioreceptors. On the other hand, specialized sensory organs are limited in their distribution and function; for example, chemoreceptors of the nasal or vomernasal organs display a range of specialized structures in the herpetofauna that may capture different modalities, such as infrared receptors of snakes like the pit vipers, than commonly presented by “higher vertebrates.” Given the typical vertebrate layout of sensory receptors and sensory organs, it is not surprising that we know relatively little regarding the effects of chemical exposure on these structures or their functions. Aquatic amphibians, like the fishes, have lateral lines to help them navigate and maintain balance. These sensory organs consist of mechanoreceptors and electroreceptors that are located within canals on the surface of the head and body. Lateral lines and their receptors help aquatic amphibians navigate, particularly when visual orientation is difficult because of murky water, and detect wave or pressure changes created by a predator’s or prey’s movement (Burggren and Just 1992; Butler and Hodos 1996). Anurans and caecilians lose their lateral line organs at metamorphic climax, while urodeles retain them (Lannoo and Smith 1989). Pheromones and specialized olfactory sensory organs are active under fossorial or other conditions of low light (e.g., in Hydromantes spp., Plethodon spp., and caecilians); auditory stimuli are used to locate calling frogs (e.g., Bufo spp.). The primary system used for detecting predator or prey in adult anurans and salamanders is visual (Brooks 1981). During the transition from water

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to air, adjustments are made in the visual system (for review of specific morphological changes, see Duellman and Trueb 1994). Focus is adjusted for changes in the fluid medium, and structural modifications are made in the cornea and lens. For example, numbers and types of photoreceptor cells increase, and there is a shift from primarily cones to primarily rods. Photopigments change from primarily 3-dihydroretinal-based porphyropsin (blue) to retinal-based rhodopsin (red), with larval pigment absorptions being at the longer wavelengths (for review, see Burggren and Just 1992; Wilczynski 1992). Eyelids are developed for life on land. As the adult anuran becomes carnivorous and its foraging mechanism changes, there is a corresponding shift to accommodate the ipsilateral optic projections, brain connections, and binocular vision needed to capture moving prey (Burggren and Just 1992). Such changes in the visual system are not limited to metamorphosis, however. Notophthalmus newts are first terrestrial, then aquatic, then terrestrial again, and pigmentation enzymes are converted accordingly. In adult anurans, random isomerization of visual receptors is temperature sensitive. Despite the relatively low body temperature of the adult frog, the threshold for receptor isomerization has an even lower set point to ensure that vision-dependent foraging and other activities are negatively affected at ambient temperatures commonly encountered during evening forays (Rand 1988; Wilczynski 1992). Very little information about chemical effects on amphibians’ vision and auditory systems has been published, but given the importance of these systems to survival, additional research is warranted. While the early literature focused on the sensory biology of reptiles remains valuable to our understanding of the interactions between organisms and their environments, there is scant work in the peer-reviewed literature regarding effects of chemical exposure on altered sensory functions, especially as relates to changes in behavior due to exposure or that might alter exposure (e.g., through contaminant avoidance). Although the interrelationships between sensory systems and behavioral responses linked to exposures to environmental chemicals have been repeatedly recognized as critical to our understanding events occurring in the field (see Grue et al. 1997; Burger 2006), responses of integrated systems such as the sensory-behavior system of the herpetofauna and their responses to environmental stressors must receive more focused study in the future.

5.7.2 Locomotion and Foraging Foraging strategies and feeding behaviors are limited by locomotion and its underlying biomechanics supported by physiology and biochemical processes. Locomotion and foraging involve a range of integrated processes reliant on interactions among gastrointestinal and nutritional physiology, neurophysiology, and skeletomuscular and behavioral mechanisms. Each of these processes may be considered at various levels of biological organization ranging from organismal to molecular, yet all reflect responses to various types of environmental stressors. Variability linked to these environmental stressors and the resulting organismal responses may provide insight into patterns of adaptation that influence exposure to environmental chemicals, since “whole animal” responses (e.g., behavioral strategies linked to foraging) to environmental conditions influence foraging, and are subsequently linked to digestive efficiency. All are intricate links to chemical exposures in the field, and outcomes of exposure may subsequently influence developmental patterns and growth, as well as other endpoints identified by ecotoxicologists.

5.7.3 Amphibians and Reptiles: Entanglements of Chemical Exposures, Foraging, and Feeding Habits If exposure to chemical stressors in the herpetofauna were dominated by dietary routes, then screening level evaluations of effects in amphibians and reptiles linked to chemical stressors would be guardedly developed under the best of circumstances. And, beyond a screening level analysis, the sparingly available existing data and previously published information clearly suggest additional

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study be considered, if herpetofauna are central to the evaluation process. While existing data suggest that amphibians sense and subsequently avoid toxins from conspecifics or predators, little, if any, information is available related to their ability to detect the presence of chemical contaminants (Steele et al. 1991). Depending on the species’ natural history and preferred habitats, chemical exposure may be reduced if avoidance behaviors are sufficient and habitat heterogeneity provides refugia to reduce or avoid chemical exposures. For example, aquatic amphibians may not be able to avoid chemical exposure if chemicals released to their local environment are widespread or their eggs or early developmental stages are limited with respect to their avoidance capabilities. Similarly, if contaminated soils are widely distributed throughout preferred habitats, reptiles and terrestrial life stages of amphibians may be chronically exposed to contaminated soils as a consequence of their high fidelity to sites within these contaminated areas (Matson 1998). Indeed, depending on the matrix involved in chemical exposure, habitats linked to critical life stages, or critical periods in a species’ life history (e.g., breeding, egg development), chemical exposures will inevitably be entangled with the integrated processes characteristic of foraging and feeding habits, as well as physiological events predicated on these starting points intended to acquire essential energy resources. Entanglements among chemical exposures, foraging, and feeding habits may ultimately yield adverse effects on integrated neurological and musculoskeletal mechansisms that diminish locomotor activity, which inevitably affects foraging. For example, in amphibians each stage of development generally has different locomotor adaptations (e.g., larval anurans have spinal cord segmentations for localized locomotor control, but these segmentations are condensed during metamorphosis), and these in part may determine differential behavioral responses consequent to exposure. After metamorphosis, locomotion in both anurans and urodeles is shifted from swimming or walking in water to traveling on land. In aquatic habitats, adult frogs swim by propelling water between their hind limbs and pushing off with the webbing between their toes, and hop by using a series of short leaps that bring both legs into the air simultaneously. In contrast, aquatic salamanders swim by undulating their tails. Most of the rhythmic activity in the anuran (i.e., tail undulation, kicking, or jumping) is due to a central pattern generator, which is an interneuronal network coordinating and synchronizing motor neurons (Burggren and Just 1992; Butler and Hodos 1996) that may be adversedly effected in response to a wide range of environmental chemicals. Hence, exposures in the field may disrupt neurological function and control of locomotor activities, which leads to deficits in foraging and predator avoidance behaviors (Rohr et al. 2003). As with any foraging or prey-predator system, feeding activities present increased risks of predation to herpetofauna. But, unless adversely affected from chemical exposure, nominal predator avoidance reduces those predation risks, as a result of integrated neurobehavioral and musculoskeletal actions yielding a range of escape behaviors. Energy reserves are required to physically avoid predators. For example, high aerobic–low anaerobic metabolism and aerobic citrate synthase are associated with slow-moving amphibians (e.g., toads), whereas high anaerobic–low aerobic metabolic adjustments and glycolytic enzyme activity are associated with quick-moving amphibians (e.g., tree frogs). As in other vertebrates, if anaerobic demands are too high, lactic acid in tissues increases, resulting in exhaustion (Pough et al. 1992). Some amphibians escape predation by secreting tetrodotoxins or by colorfully advertising their toxin production, but scant information on effects of chemical exposure related to color changes or toxin secretions has been reported. Similarly, little, if any, characterization of chemical effects on integrated functions critical to foraging and acquisition of energy stores is available for reptiles. Comparative physiologists and physiological ecologists strive to synthesize generalizations from the diversity of traits observed in animals, particularly as those generalizations relate to integrated functions such as locomotion and derivative activities such as foraging. Variations due to size, temperature, locomotor mechanisms, gaits, differential influence of Reynolds number and lotic habitats on body size, and capacity to store mechanical strain energy all contribute to this variation, and energy resources serving these functions display a similar range in diversity. Regardless of

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their preferred foods and the time course for acquisition, foraging incurs a major energetic expense across the wide range of species of herpetofauna (reviewed in Pough et al. 1992; Bridges 1997). These energetic costs depend on environmental factors such as distribution of prey and predators, temperature, habitat quality, pH, and humidity, each of which may interact with chemicals in the environment. Environmental chemicals directly and indirectly affect an animal’s foraging and feeding activities, or the assimilation of foods once consumed, for example, by reducing or eliminating forage or prey items available in the field, by incorporation of tissue residues into the food base and rendering it toxic, and by altering the animal’s ability to consume and assimilate the food (Hall 1990; Walker et al. 1996). Risk of exposure to environmental chemicals and their effects will depend, in part, on the animal’s foraging behaviors and feeding strategies. As noted in Section 5.3.1 some amphibians and reptiles can survive without food for extended periods, ranging from weeks to months to years. In other species, risks linked to foraging behaviors and feeding strategies will also vary as a function of developmental stage. For example, during metamorphosis in amphibians, feeding is reduced, growth becomes arrested, and developmental events characteristic of metamorphosis rely on available energy stores, including those derived from tail resorption. During metamorphosis, changes in diet are associated with changes in the digestive structures. The foregut-midgut of herbivorous anuran larva stores food and associated glands help in extracellular digestion. Absorption is maximized by the characteristic gut coiling and extensive production of pancreatic and hepatic enzymes. As the gut regresses, the coiling is lost, and pepsin-secreting cells form the functioning stomach, while larval laminar cilia are replaced by functional microvilli (Duellman and Trueb 1994). Parallel to these changes, the pancreas is restructured as a functional component of the endocrine system, and kidney and liver enzyme functions shift from excreting water to conserving fluids, and from producing ammonia to producing urea waste (Duellman and Trueb 1994). With the change in available oxygen, hemoglobin production and control by the spleen and liver are increased. The integrated outcome yields postmetamorphosic anurans as carnivores. Salamanders do not undergo such extreme morphological transitions and are therefore able to feed indiscriminately on aquatic invertebrates and algae and, as they grow, other larval amphibians. Throughout the herpetofauna, foraging tactics vary across species. Within a species, foraging tactics may differ from habitat to habitat, and within-habitat differences in predatory and foraging behaviors will depend on prey or forage items available (e.g., foraging in preferred habitat may differ from that displayed in marginal habitats). In both amphibians and reptiles, selection of forage and prey items may change with increased animal size (Leff and Bachmann 1986); for example, maximum prey size may increase as a function of gape size, which in turn may reflect the changing energetic demands linked to growth and reproduction. Effects of chemicals occurring in food items may affect growth of early life stages, and rates of development may be critical in determining whether the organism will meet its nutritional requirement as it matures. Growth and associated gape size also influence consumption rates, which inevitably influence ingestion of environmental chemicals contained in food sources. Regardless of the role that chemical stressors play as components in an animal’s interaction with the environment, a general metric for success of individuals or a population is the balance between energetic costs and benefits directly or indirectly related to these costs (see Cohen 1978 and Stephens and Krebs 1986). For example, foraging by terrestrial herpetofauna reflects a dynamic balance involving a wide range of behaviors and “hunting” techniques (regardless of their being carnivores or herbivores) that vary in duration, locomotor costs, and energetic rewards. Various feeding strategies are reflected in the life histories of amphibians and reptiles, and may conveniently be categorized based on feeding habits. The adaptive interplay between foraging and feeding behaviors, and the capacity to regulate digestive performance among amphibians and reptiles (as well as vertebrates in general) clearly have implications for evaluating exposures to environmental chemicals. While recent activity indicates an increased awareness of research needs to characterize endpoints pertinent to amphibians and reptiles, much additional work must be

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completed to achieve parity with other vertebrates when topics related to chemical exposure and effects are considered.

5.8 Biomarkers, Metabolism, and Development of Energetics-Based Tools Biomarkers have increasingly been applied to field and laboratory studies focused on chemical exposures and effects in herpetofauna. For example, Venturino et al. (2003) identified biomarkers commonly applied to studies focused on amphibians, and Linder, Lehman, and Bidwell (Chapter 4, this volume) noted that biomarkers have increasingly been applied to evaluating exposure and effects in herpetofauna exposed to environmental chemicals in aquatic habitats. Walker et al. (2001) characterized biomarkers as morphological alterations, genetic effects, behavioral parameters and tissue residue levels, which extended Huggett et al. (1992), who had identified biomarkers as biochemical, physiological, and histological endpoints used to evaluate exposures and effects of chemical stressors. Biomarkers used in ecotoxicological investigations focused on amphibians and reptiles are similar, if not identical, to those applied to studies of other vertebrates. Ideally, by using biochemical and physiological endpoints to evaluate exposure, adverse effects linked to chemical exposures may be anticipated before responses are observed at organismal or population levels of organization (Newman and Unger 2003; Mitchelmore et al. 2006). Although data may not be available for evaluating the status of herpetofauna across their current range of habitats (e.g., comparing populations presumptively exposed in the field to similarly collected data for populations at reference areas; see Henry 2000), long-term acquisition of such data will undoubtedly contribute to future characterizations of chemical exposure and the general health status of the herpetofauna existing under a wide range of environmental conditions in the field (Rie et al. 2000; Venturino et al. 2003). While larger-bodied herpetofauna (e.g., turtles) have historically found wide use in studies concerned with measurement of tissue residues, small body size and limited blood volume characteristic of many species of amphibians and reptiles may account for the lack of reference or baseline data for biochemical and hematological attributes. These measures vary with factors such as species, developmental stage, gender, reproductive status, season, and nonspecific stressors (Zhukova 1987). In contrast to data available for fishes, few reference data for biochemical and physiological markers (e.g., hepatic oxidized/reduced glutathione ratios, lipid peroxidation, other tissue-specific indicators, and routinely measured serum or plasma chemistries) have been compiled for herpetofauna. Although designed laboratory studies characterizing baseline conditions are encouraged to offset data deficiencies, opportunistically recording these data may help us to accumulate baseline measures and decrease the variance for future reference (Henry 2000). Biochemical and physiological biomarkers for the herpetofauna include endpoints shared with other vertebrates. For example, in studying exposure and effects of lead in amphibians, Loumbourdis (2003) considered histological effects — the development of kidney inclusion bodies — in Rana ridibunda, while Arrieta et al. (2004) focused on the disruption of heme synthesis by measuring intermediate metabolites or degradation products (e.g., porphyrins) and altered aminolevulinic acid dehydratase (ALAD) activities in lead-exposed Bufo arenarum. Other studies applying biomarkers to the analysis of exposure and effects of environmental chemicals in herpetofauna have focused on enzymes frequently used in studies of other vertebrates exposed to chemicals in the field (e.g., Sparling et al. 2001, van den Brink et al. 2003 on amphibians, and Clark et al. 2000 on reptiles). Evaluations of enzyme activities characteristic of xenobiotic metabolism of polynuclear aromatic hydrocarbons and organochlorine compounds in amphibians and reptiles (e.g., measurements of mixed-function oxidases and associated enzymes) have been reported (Venturino et al. 2001; Gunderson et al. 2004; Kostaropoulos et al. 2005). For example, induction of liver enzymes that are part of the mixed-function oxidase (MFO) system has been used to indicate exposure to a range of organic chemicals (including pesticides) in a number of vertebrates, but reduced activity of these

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enzymes in amphibians may limit their use as a biomarker for this group (DeGarady and Halbrook 2003; Venturino et al. 2003). While components of the MFO system are present in amphibians and reptiles, their activities occur at levels lower than those measured in mammalian systems (Schwen and Mannering 1982a, 1982b, 1982c; Ertl and Winston 1998). Desulfuration, hydroxylation, epoxidation, conjugation, reduction, and hydrolsis reactions are readily measured in tissues (e.g., liver), but induction of cytochrome P450 is consistently lower in the herpetofauna than in birds and mammals (Ertl and Winston 1998; Huang et al. 1998), which may reduce its effectiveness as a biomarker of exposure. The capacity to detoxify chemicals via MFOs will vary from substrate to substrate and will likely differ for aquatic and terrestrial amphibians. Similarly, MFO activity will vary as a function of developmental stage. For example, Rana catesbeiana tadpoles collected from chemically polluted sites exhibited higher metabolic rates than did those collected from reference sites (Beatty et al. 1976; Rowe et al. 1998), yet developmental time may influence the capacity for enzyme induction across a range of species. In general, species of herpetofauna may have developed mechanisms of resistance to environmental chemicals as a result of natural selection (Boyd et al. 1963; Hall and Kolbe 1980). Studies focused on herpetofauna and the role that environmental chemicals have in disrupting endocrine function illustrate recent efforts to measure biomarkers of exposure and effects in field and laboratory studies. For example, numerous studies have considered exposure to endocrinedisrupting chemicals and subsequent effects on gonad morphology, sex hormones, and the levels of reproductive hormones in various amphibians and reptiles (see, e.g., Hayes et al. 2002, 2003; Yang et al. 2005), which reflect increased research beyond Hayes (2000) and Guilette (2000). Indeed, amphibians and reptiles have played an important role as indicators of EDCs in aquatic systems. Most work to date has focused on effects related to sex determination, with biomarkers including gonadal morphology or circulating levels of plasma hormones or specific proteins (e.g., Noriega and Hayes 2000; Shelby and Mendonca 2001; Hayes et al. 2003). Many egg-laying reptiles may be particularly suited for studies of chemicals that affect sex hormone balance, since they normally have TSD, which facilitates manipulation of sex ratios to more clearly evaluate chemical effects (Crews et al. 1995; Newman and Unger 2003). Another biochemical marker frequently applied to studies focused on herpetofauna exposed to EDCs is vitellogenin (see Section 5.4.3.1). Although males and females are both genetically capable of synthesizing vitellogenin, production is induced by estrogenic compounds; hence, vitellogenin is a relatively sensitive biomarker of estrogenic chemical exposure to males (Palmer et al. 1998). Similarly, recent studies also indicate a significant role for herpetofauna in the evaluation of chemicals that disrupt the thyroid axis. Effects on thyroid hormones — thyroxine (T4) and triiodothyronine (T3) — linked to exposures to environmental chemicals have been observed in a range of amphibians and reptiles over the past 10 to 15 years (Gunderson et al. 2002; Tada et al. 2004; Mosconi et al. 2005; Yang et al. 2005). The role of thyroid hormones for initiating hatching and the onset of metamorphosis was well studied in herpetofauna prior to the ecotoxicologists’ focus on endocrine-disrupting chemicals released to the environment (Brasfield et al. 2004; Furlow and Neff 2006; Tata 2006). Given the heightened awareness of emerging contaminants that occur in treated effluents and water treatment residuals such as biosolids, future work focused on effects linked to altered endocrine function will inevitably increase. As in birds and mammals, metallothionein1 (MT) has been used as a biomarker for exposure to metals, such as copper, zinc, and cadium, which induce MT synthesis in herpetofauna as they do in other vertebrates (Suzuki and Akitomi 1983; Vogiatzis and Loumbourdis 1998). Similarly, as a diagnostic tool for exposure to organophosphorus (OP) and carbamate pesticides, cholinesterase 1

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Metallothioneins (MTs) are low-molecular-weight proteins found in all eukaryotes (often in multiple copies) as well as some prokaryotes. MTs are unusually rich in cysteine residues that coordinate multiple zinc and copper atoms under physiological conditions. Cadmium-induced synthesis of MTs has been observed in herpetofauna, as was previously described in fishes and wildlife.

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inhibition measured in serum, plasma, red blood cells, or other tissues (Sparling et al. 2001) has been increasingly applied in ecotoxicological studies focused on amphibians and reptiles. Additional studies relying on measurements of cholinesterase activity as a tool to evaluate OP and carbamate exposure should be encouraged to ensure the tool’s value being comparable to that seen in other vertebrate classes (Baker 1985; Rosenbaum et al. 1988; Bonin et al. 1997). Research on amphibian response to OP and carbamate chemical effects indicates a wide range of species sensitivity as measured by various endpoints (Hall and Kolbe 1980; Rosenbaum et al. 1988; Snawder and Chambers 1993; Honrubia et al. 1993; Sparling et al. 1997; Taylor et al. 1999a, 1999b, 1999c, 1999d). As with other vertebrates, field and laboratory investigators have also employed a range of diagnostic tools to evaluate exposure and effects, including clinical chemistry analyses on serum and tissue samples (see, e.g., Papadimitriou and Loumbourdis 2005). Relatively new to ecotoxicology and herpetofauna research are immunological markers, including tools that evaluate cellular immune function through nonspecific cytotoxic cells and macrophages. These tools have been advanced by researchers evaluating a variety of chemical stressors (e.g., R. pipiens exposed to low pH; see Vatnick et al. 2006). At present, relatively little is known about the immune system in most of the species of amphibians (Carey and Bryant 1995; Taylor 1998), and based on the differential susceptibilities to infections such as red leg caused by Aeromonas hydrophilla, there are potentially significant interspecific differences. Amphibians possess the major tissues associated with immune response, such as thymus, spleen, kidney, bone marrow, and lymphoid cells (Carey and Bryant 1995), as well as the ability to induce antibody response to initial and subsequent antigen exposure. There are differences in immune response depending on stage of development, and the immune system seems to undergo near-complete change during metamorphosis (Carey and Bryant 1995). Immune cell function, however, is also temperature and season dependent (for review, see Taylor 1998). Although comparative immunologists have considered the herpetofauna over many years, ecotoxicologists have yet to benefit from that experience, in many respects tracking the history of the discipline’s experience with fishes, birds, and mammals. Matching the increased interest in immunological markers are tools focused on measurement of genetic markers. Expanded suites of tools focused on genetic markers have been developed in the recent past focused primarily on vertebrates, but these tools are relatively underemployed in studies targeted on herpetofauna. Measures related to DNA strain breakage and sister chromatid exchange have been applied to studies of herpetofauna (see, e.g., Wirz et al. 2005; Tverdy et al. 2005), but the area is potentially rich for development by ecotoxicologists. Various tools are available in the physiological ecologist’s tool box that would be amenable to application to ecotoxicological studies, including energetics-based biomarkers that could provide a common currency — energy and materials — that directly links individual, population, and community levels of organization (Congdon et al. 2001). Rowe et al. (2003) discussed energetics as it relates to larval, juvenile, and adult stages of anuran amphibians, and clearly identified the role that chemical stressors may play in increasing maintenance costs and decreasing energy available for growth. Along similar lines of discussion, reptilian eggs may also serve as valuable models to study the energetic effects of chemical stressors, since development of the embryo relies entirely on internal yolk stores, and contaminants may pass across the eggshell in association with imbibed water (Moeller 2004). Application of energetics analysis has been advocated for studies focused on amphibians and reptiles (Rowe et al. 2003), and a wide range of tools are potentially available to the ecotoxicologist. For example, 1 tool commonly deployed to study energetics in vertebrates — measurement of specific dynamic action — has received only limited use by ecotoxicologists in their study of herpetofauna.

5.8.1 Specific Dynamic Action Specific dynamic action (SDA) represents the summed energy expended on ingestion, digestion, and assimilation of food (Brody 1945; Kleiber 1975). Given likely directions of regulatory applications

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of food chain analysis as a tool to evaluate risks, deploying SDA in conjunction with integrated field and laboratory studies focused on dietary exposures to environmental chemicals seems pertinent to developing our tool box for evaluations of herpetofauna. SDA directly relates to the presumptive role of diet as a critical link between environmental chemicals and receptors. Before an animal can allocate ingested energy to growth, maintenance costs supporting daily functions and metabolism must be attained (Angilletta 2001). Physiological processes that contribute to SDA include gastrointestinal motility, production of digestive enzymes and nitrogenous wastes, protein catabolism and synthesis, and intestinal nutrient transport (Jobling 1981; Hailey 1998; Secor 2003; McCue 2006). Variations in the relationship between SDA and nutritient content of prey can influence growth, which can in turn impact survivorship, reproductive success, and ultimately fitness (Brodmann et al. 1997; Rosen and Trites 2000; Babu 2001). SDA reflects maintenance costs associated with food processing and, depending on species, has a varying impact on the net assimilated energy available for growth and reproduction, endpoints commonly measured in ecotoxicological studies. In contrast to endotherms (see Costa and Kooyman 1984 and Hawkins et al. 1997), for amphibians and reptiles, the process of ingestion, digestion, and assimilation of food accounts for a much greater increase in metabolic responses (e.g., increased metabolic rates; see Coulson and Hernandez 1979; Secor and Diamond 1997a, 1997b; Secor and Phillips 1997; Powell et al. 1999; Secor 2001). Furthermore, these marked increases in metabolism are captured by an immedidate postprandial metabolic response and an extended period beyond the postprandial response in which metabolic rates are increased during the digestion and assimilation process (e.g., at least 2 weeks; see Secor and Diamond 1997a, 1997b; Secor 2005a, 2005b). As a consequence, from the perspective of developing an energy budget for a given animal, SDA contributes significantly to an ectotherm’s energy budget (Secor and Nagy 1994; Peterson et al. 1998). While work with SDA has considered a wide range of metabolic responses experienced by amphibians and reptiles during digestion and assimilation, only limited consideration has been given to toxicant interactions with quality and quantity of meal, feeding frequency, body temperature, or body size as that tracks age or availability of food sources (see Secor 2005a, 2005b; McCue 2006; Secor and Boehm 2006). SDA as a “measure of effects” would enable ecotoxicologists to tap into existing literature in the physiological ecology of amphibians and reptiles. As such, using SDA as a measurement endpoint might benefit the evaluation of chemical effects linked to multiple stressor exposures in the field. Studies focused on measurement of SDA have relied heavily on amphibians and reptiles as experimental models, which clearly suggest that SDA and other tools of the physiological ecologist may be applied by ecotoxicologists to address toxicant effects in animals presumptively exposed predominantly via diet. Overall, the role of biomarkers in evaluating herpetofauna exposure and effects and differences in biochemical and physiological characteristics between animal groups are presently incompletely understood, which may initially affect the utility of some variables for indicating contaminant exposure in regulatory applications. Nonetheless, our current implementation of biomarkers for the study of herpetofauna exposed to chemicals in the field is better developed than 10 years ago. Research over the next 10 years should refine those tools to a greater extent and allow time to more adequately develop tools potentially beneficial to the evaluation process for amphibians and reptiles.

5.9 Interactions of Chemicals with Physiological and Environmental Factors Exposures in the field and multiple stressors are commonly linked, oftentimes intractably, which contributes to confounded interpretation of effects associated with exposure to chemical stressors. Interactions between a wide range of variably responsive receptors and environmental factors, including chemical stressors, define the common ground of an ecotoxicologist and a physiological ecologist (see Relyea, Chapter 14, this volume). There are innumerable ways in which environmental factors can interact with the physiology of the herpetofauna. Here we illustrate 2 commonly

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encountered environmental factors — ultraviolet-B (UV-B) radiation and stress — that are repeatedly encountered in characterizing effects associated with chemical exposures.

5.9.1 Ultraviolet Radiation Ultraviolet (UV) radiation is that portion of the electromagnetic spectrum occurring between x-rays and visible light, having wavelengths between 40 and 400 nm. Its spectrum has been divided into vacuum UV (40 to 190 nm), far UV (190 to 220 nm), UV-C (220 to 290 nm), UV-B (290 to 320 nm), and UV-A (320 to 400 nm; see ISO/DIS 21348 [ISO 2005] for additional specifications on characterization of UV radiation). From an ecotoxicologist perspective, however, work focused on UV radiation relates primarily to observations of adverse effects linked to UV-B exposures, particularly those associated with elevated fluxes of UV-B radiation linked to ozone depletion. In applications to human and veterinary health, UV-B has long been studied in animal models because of its role in the synthesis of vitamin D3, since UV-B at wavelengths between 270 and 300 nm (peak synthesis occurs between 295 and 297 nm) initiates conversion of 7-dehydrocholesterol to vitamin D3. Historically, pathological and toxicological studies on UV-B have focused on adverse effects linked to prolonged exposures to nominal atmospheric fluxes, yet recent findings of diminished atmospheric ozone and the resulting increase in UV-B point toward singly or jointly acting effects linked to UV-B exposure. In terrestrial and aquatic systems, UV-B is the wavelength of UV radiation of primary concern. In aquatic systems, UV-B penetration into freshwaters is strongly influenced by altitude, by the extent of plant canopy adjacent to the habitat, and by the concentration of dissolved organic material (DOM) in the water. DOM in surface waters results from heterogeneous inputs of decomposing plant, microbial, and animal materials that act as the primary absorber of UV-B. Interactions between DOM and UV-B may lead to photodegradation of these organic materials (Häder et al. 1998; Häder 2006), which in turn could promote positive feedback wherein UV-B exposure leads to greater UV-B flux and less DOM protection. In field settings, plant canopy characteristic of adjacent terrestrial and wetland habitats influences input of UV-B and other wavelengths of radiation to aquatic systems by absorbing and reflecting UV-B. Depending on the vegetation and the incidence of radiation, plant canopy may remove up to 90% of the incident light (Xenopoulos and Schlinder 2001). Herpetofauna cannot help but be affected by increased incidence of UV radiation across the habitats upon which they depend. Little and Calfee (Chapter 13, this volume) extend previous reviews focused on UV radiation and its effects on freshwater vertebrates (Little and Fabacher 2003). In part, these reviews coincidentally followed from observations of Henry (2000) that depletion of atmospheric ozone significantly influenced the increase of UV-B globally and that increased incidence of UV-B inevitably played directly as a physical stressor for a wide range of receptors, including members of the herpetofauna. Interactions between UV-B radiation and chemical stressors have been increasingly reported as jointly acting physical-chemical stressors that serve reactive chemical species in the exposure mileau of a wide range of receptors (Little and Calfee, Chapter 13, this volume; Little and Fabacher 2003). While photodegradation contributes to fate processes for chemicals in the environment, UV-B photoactivates chemicals such as polynuclear aromatic hydrocarbons (PAHs) to yield hydroxyl radicals and oxides (see Sparling, Chapter 9, this volume). In part, the complexity of exposure in the field and the “thrust and parry” exchanges between stressors and receptors are captured by entangled systems that involve UV-B photoactivation, chemical stressors, and biota, since an animal’s susceptibility to the UV-B effects may depend on its metabolic ability to bind and clear these highly reactive compounds (Ovaska 1997; Walker et al. 1998). Species-specific sensitivity has been correlated to photolyase activities and to specific life history strategies. Photolyase provides a repair mechanism for the DNA molecule damaged by radiation (Ovaska 1997). The enzyme is present and active in the less chemically

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sensitive species (e.g., Rana aurora, Hyla regilla) and is less active, if present at all, in the more sensitive species (e.g., Rana cascadae, Bufo boreas, Ambystoma gracile; Blaustein et al. 1996). Another major factor determining sensitivity to UV-B is the individual’s likelihood of being exposed to UV-B. Radiation effects are cumulative; therefore, the thickness and moisture of the individual’s skin surface and the burrowing habits and daily and seasonal activity of the individual are important considerations. DNA damage is particularly deleterious during embryonic development or metamorphosis, and factors such as egg pigmentation, distribution and arrangement of eggs and egg masses (laid singly, as sheets, or in masses), and the depth and clarity of the water in which eggs are deposited should also be considered. For example, under reference conditions, Bufo bufo eggs deposited in deep water are less exposed to sunlight than are the eggs of B. calamita, which are deposited surfically. If environmental conditions change, such as during global warming or drought, the B. bufo embryos with less photolyase could become more susceptible to UV-B effects than embryos developed from B. calamita (Lizana and Pedraza 1998). These species-specific sensitivities may be linked to population level effects, given studies on anurans and salamanders completed in the recent past that suggested that ambient levels of UV-B (290 to 320 nm) can affect individuals and populations (Blaustein et al. 1994b, 1996; Lizana and Pedraza 1998). Direct UV-B effects observed include embryonic mortality and failure to hatch (Blaustein et al. 1994a, 1996), abnormal larval development, increased limb and musculature deformities, neurological damage, immunosuppression, and increased cellular damage in the eyes and skin surface (see Ovaska 1997 for review). Indirect effects include increased susceptibility to fungal infestations (Lizana and Pedraza 1998). A biochemical assay to measure photolyase activity in oocytes is available (Blaustein et al. 1994c) and helps to evaluate the ability of a species to repair DNA following UV-photoinduced damage. Such a marker provides a measure of whether a species may be at increased risk; additional information on the exposure to UV-B based on behaviors helps to complete the evaluation. In terrestrial systems, amphibians and reptiles are exposed to altered fluxes of UV-B and other atmospheric gases, and responses to these fluxes will range widely in a dose-responsive manner, ranging from inconsequential to beneficial effects through extinction level events linked to elevated fluxes of UV and other deleterious electromagnetic radiation (see Cockell 1999). In herpetofauna, UV-B plays roles similar to those observed in typical mammalian and avian models, wherein UV-B stimulates synthesis of vitamin D3, which ultimately follows various speciesspecific metabolic pathways characterized by a range of biological activities. For example, fish, amphibians, reptiles, and birds rely on vitamin D3, while many mammals benefit from vitamin D3 or vitamin D2. Although the literature detailing UV-B effects on reptiles is increasing relative to the literature available a decade ago, it remains sparse. However, there are no reasons to discount exposure of reptiles to elevated UV-B more than other vertebrates, given the predisposing behaviors of many reptile species. For example, terrestrial saurian reptiles frequently depend upon sun basking, sunshade shuttling, and other heliothermic behaviors for regulation of core body temperature; these behaviors will ensure that these animals experience potentially significant exposure to solar UV radiation (UVR). These exposures may place them at increased risk of deleterious UV-B effects, especially given the absence of protective feathers or pelage common to other terrestrial vertebrates. Although the keratinized skin and scales of reptiles may confer protective benefits that offset increased incidence of UVR, given the range of effects that have been exhibited by other terrestrial vertebrates subjected to increased UV-B exposures, reptiles are likely to present similar responses when exposed. For example, cutaneous UV-B exposure alters immune function in rodents, for example, inhibition of delayed-type hypersensitivity reactions (Kim et al. 1998) and splenic and peritoneal macrophage functions (Jeevan et al. 1995). In fish, effects have been observed in laboratory exposures involving single acute low-dose UVR exposures of non-UVR-adapted fish. UVR exposures yielded stimulation of whole blood phagocyte respiration, but demonstrated a decreased activity of head kidney granulocytes (Salo et al. 2000). Observations have also been recorded in fish

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that suggest species tolerant of UV-B radiation contain a methanol-extractable nonmelanic photoprotective substance in the skin (Blazer et al. 1997; Fabacher and Little 1995, 1998). In contrast to these findings for fish, the heliothermic green anole appears resistant to UVR inhibitory effects on cutaneous cell-mediated immune responses and splenic phagocytic function. These observations in reptiles may result from a combination of epidermal-dermal factors and are not linked to extractable photoprotective substances synthesized in the skin (Cope et al. 2001). Currently, mechanisms associated with UVR resistance are incompletely characterized in reptiles, and development of solar UVR-induced immunosuppression in green anoles and other members of the class Reptilia should not be underestimated. Our understanding of humoral and cellular immune responses to cutaneous UVR exposure is largely unknown, and given their phylogenetic distance, UVR-induced immunosuppression in reptiles may be markedly different from those mechanisms characterized for mammals. Given observations that amphibians and reptiles may be expressing an increased susceptibility to disease in the field, immune responses such as delayed-type hypersensitivity reactions and systemic macrophage functions may be useful markers of UVR effects on the reptilian immune system.

5.9.2 Stress Whether acting singly or interactively, changes in environmental conditions shape responses in systems at various levels of biological organization. These environmental stresses (e.g., temperature, ambient radiation and other physical stressors, and chemical and biological stressors) acting jointly with environmental chemicals can potentially work in various ways to stress an animal. For example, from an energetics perspective, chemical stressors may directly or indirectly disrupt energy balance either by decreasing the resources available or by increasing the energy required for maintenance (e.g., eliminating insect prey through pesticide use will decrease energy sources). Chemical stressors acting singly or jointly with UV-B may physically damage epithelium and predispose the animal to diseases, bacterial infection (Faeh et al. 1998), pathogenic fungi (Taylor et al. 1999a, 1999c, 1999d), or water mold (Lefcort et al. 1997). Environmental chemicals also act as nondistinct stressors and become part of the sublethal environmental changes (e.g., EDCs or other chemically induced problems in steroid feedback, metabolic activity, sensory organ function, gaseous exchange, or liver function). As a stressor, toxicants may activate an organism’s normal stress response (corticosterone release) and eventually compromise their ability to respond to stress (Gendron et al. 1997; Hayes et al. 1997; Hopkins et al. 1997). Despite their many adaptations, amphibians and reptiles are susceptible to synergistic or additive effects of multiple stressors. From the perspective of the physiological ecologist, generalized responses to stress are readily apparent and well characterized, but mechanisms linked to these organismal responses continue to be objects of research across a wide range of animals. Generalized responses to stress, oftentimes linked to extremes in environmental temperature, resource availability (e.g., seasonal variations in prey or vegetation), or limited water, may serve as existing physiological adaptations to offset exposures to environmental chemicals. Dormancy is a commonly observed response to unfavorable environmental conditions, conditions potentially characterized by the occurrence of stressors that exceed species-specific preferences. For example, dependence upon external conditions for regulating metabolic rate limits the distribution of amphibians and reptiles, and when prevailing conditions exceed species-specific tolerances, animals will reduce activity and enter dormancy until acceptable environmental conditions return. In most temperate species, periods of dormancy are a normal feature of their yearly cycle of activity, and are most often adaptations to avoid or minimize exposures to seasonal extremes in temperature or moisture levels. For some species, dormancy may account for a significant portion of their yearly cycle. Dormancy takes 2 major forms: hibernation for the avoidance of cold and estivation for avoidance of other environmental factors, such as drought (Gregory 1982).

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5.9.2.1 Hibernation Anticipatory or seasonal prehibernation adjustments in response to cold prepare the animal for the depletion of energy stores and restriction of caloric intake (Herman 1992; Pinder et al. 1992). Homeoviscous acclimation refers to biochemical and structural changes made at the level of the membranes in response to colder conditions. Such modifications may result in a greater percentage of unsaturated fatty acids, an increase in permeability, and control of transport mechanisms (Pinder et al. 1992; Crockett 1998). Food is converted to glycogen and lipids, is stored in the liver and fat bodies, and metabolism is depressed. The extent to which stores are used or conserved depends on whether the amphibian is terrestrial or aquatic (e.g., toads depend on lipids, frogs on glycogen), their prior thermal acclimation (e.g., species living at higher altitudes or in colder climates can avert starvation longer), and the degree to which they may be freeze tolerant. Submerged aquatic amphibians risk anoxia and osmotic stress but are well hydrated. Laboratory studies indicate that, within limits, anoxic, cold submerged frogs can maintain cellular adenosine triphosphate (ATP) by increasing their carbohydrate metabolism, by using muscle, liver, and heart glycogen stores, and by depressing metabolism (Donohoe and Boutilier 1998). Terrestrial amphibians in hibernacula risk dehydrating, freezing, and accumulating toxic nitrogenous wastes, but they are safe from predators and generally do not lack oxygen. They hibernate below the frost line to avoid sudden freeze, whereas aquatic amphibians will move deeper into the ponds to reduce the risk of surface freeze. In temperate species, winter presents a significant physiological challenge. This challenge is met by both behavioral and physiological means. Hibernation can be divided into 4 stages: fasting, entering the hibernaculum, dormancy, and metabolic depression (Gregory 1982). Decreasing temperature and light are generally regarded as stimuli for entering hibernation. Declining temperatures may also suppress appetite (Gatten 1974) and initiate the fasting associated with hibernation. As temperatures fall, many species will seek refugia, or hibernacula, that will not freeze during the coming winter months. During hibernation, reptiles depress many physiological processes because food and even air may be inaccessible. To survive for several months, metabolic functions are slowed during hibernation, even more than predicted by the decreased body temperature (Zug et al. 2001). This indicates that some physiological processes have been curtailed. This reduced physiological state can conserve valuable energy supplies for significant periods of time. When metabolism is slowed, breathing and heart rate are reduced, but the supply of blood and oxygen to vital organs is maintained to ensure survival. Aquatic species may hibernate underwater beneath a layer of ice. Most of the water below the ice in lakes and streams will not drop below 4 °C. However, because the surface layer of ice prevents access to air, normal pulmonary respiration must be curtailed for these air-breathing animals. Some reptiles (e.g., Chrysemys picta, Sternotherus odoratus, and Thamnophis sirtalis) maintain aerobic metabolism from cutaneous respiration (Zug et al. 2001). However, because of the thickly cornified skin of reptiles, buccopharyngeal or even cloacal respiration may be required to maintain adequate oxygen levels for aerobic metabolism (Seymour 1982). Some aquatic turtles may burrow into the mud at the bottom of a lake or stream, preventing access to oxygenated water. The anoxic or hypoxic environment caused by burrowing in the mud leads to prolonged periods of anaerobic metabolism. However, even these turtles may shuttle back and forth from the mud to open water, where they can switch to aerobic metabolism and flush their system of the accumulated lactic acid (Zug et al. 2001). For terrestrial hibernators, physiological demands may not be as great, although even they must find shelter from freezing conditions. Typically, this means burrowing below the frost line. However, complete inactivity may not be possible. As the frost line descends, box turtles (Terrapene carolina) have been observed to burrow deeper (up to 0.5 m) to avoid freezing (Legler 1960). Hibernating snakes (Elaphe spp., Crotalus spp.) move to remain in the warmest part of their den or crevice (Zug et al. 2001). Intestinal response to long-term aphagia has been studied for infrequently feeding

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snakes (Secor and Diamond 2000) and estivating amphibians (Secor and Diamond 1996). In each case, the intestine downregulates by changing function, morphology, or both to some extent to decrease performance. Amphibians and reptiles that inhabit temperate regions of the world hibernate, which like estivation may be characterized by extended periods of aphagia (Gregory 1982; Pinder et al. 1992). Since many of these temperate species feed frequently during the summer, they would be expected to narrowly regulate digestive performance during that time. However, the influence of these functional and structural responses on exposure is poorly understood and, more critically, greatly undervalued. 5.9.2.2 Freeze Tolerance Freezing is lethal to most reptiles because the formation of ice crystals causes the lysis of cells. However, many temperate reptiles can withstand brief periods of supercooling (1 to 2 °C) in which ice crystallization does not occur (Lowe et al. 1971; Claussen et al. 1990; Claussen and Zani 1991; Packard and Packard 1995). Freezing of extracellular fluids results in dehydration. Because ice crystals form from pure water first, ions and other dissolved substances are excluded, raising the osmotic potential of the remaining fluids. This causes an osmotic imbalance, leading to dehydration of the surrounding cells and tissues. Freezing of extracellular fluids also interrupts blood and lymph flow and blocks transport of oxygen, CO2, nutrients, and waste products. Some reptiles have evolved physiological mechanisms to help them cope with freezing conditions. In some species, glucose is mobilized to act as a cryoprotectant, inhibiting freeze damage to the cells (Storey 1990). Water also may be redistributed from the tissues into the coelomic and subdermal spaces (Costanzo et al. 1993). By minimizing the amount of water in the tissues, damage from freezing can be reduced (Lee et al. 1990, 1992). However, a few species of reptiles (e.g., T. carolina, C. picta, and Alligator mississippiensis) are tolerant of some extracellular freezing (Hagan et al. 1983; Costanzo 1988; Storey et al. 1988; Storey 1990; Costanzo and Lee 1990; Costanzo et al. 1993; Packard et al. 1993). 5.9.2.3 Estivation Unlike hibernation, estivation is associated with other environmental factors besides low-temperature avoidance and is usually correlated with water conservation (Espinoza and Tracy 1997; Storey 2002) or high-temperature avoidance (Lambert 1993; Bayoff 1995; Storey 2002). Estivation occurs predominantly in turtles and squamates in hot, arid environments where they retreat into shelters deep enough to avoid excessive heat and extreme temperature fluctuations (Voigt and Johnson 1976). During estivation, reptiles exhibit a reduced metabolic response to temperature (Abe 1995), but metabolic processes are not curtailed as significantly as during hibernation. Cellular mechanisms responsible for the metabolic torpor are similar to those during hibernation (Mauro and Isaacks 1989). Temperature clearly has important implications for reptilian life cycles. Thermal pollution, therefore, may have a significant impact on reptiles by affecting metabolic rates and energy balances. This is particularly important for animals on restricted energy budgets, such as during hibernation or estivation. In addition, thermal pollution may have devastating impacts on the development of embryos in species that exhibit TSD.

5.10 Physiological Ecology and Multiple Stressors: Developing a Common Currency to Evaluate Chemical Exposures to Amphibians and Reptiles in Field Settings Sparling et al. (2000a) clearly characterized existing data and applied research needs for the ecotoxicologists encountering amphibians and reptiles in the field. Although ecotoxicology had developed a process for evaluating chemical exposures in a few species of fish and wildlife, Sparling

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et al. (2000a, 2000b) emphasized that the available literature through 1998 was, at best, sparse compared to publications focused on fishes and terrestrial vertebrates (particularly birds and mammals). Although awareness of herpetofauna has increased in the intervening period since publication of that first edition, we find a common refrain in this second edition (see Chapter 1, this volume); the literature focused on the herpetofauna again lags behind that for birds and mammals. The past 8 to 10 years have yielded a dramatic increase in research on amphibians and reptiles and the effects of environmental chemicals on these animals, yet much work remains to be conducted, if the herpetofauna are going to be sufficiently represented in the environmental risk assessment process. Few studies have characterized the role of nutritional or energetic interactions that influence exposure and mediate biological effects in amphibians and reptiles, be those direct, collateral, or indirect effects. Biological factors related to nutritional and bioenergetic interactions that influence exposure are generally not measured in screening level evaluations of chemical risks to fish and wildlife. However, if these factors are estimated or, better yet, measured, then exposures in the field might be better characterized, especially those involving long-term, low-concentration exposures. Furthermore, if data gaps were addressed using integrated field and laboratory studies (see, e.g., Linder et al. 1991; Sadinski and Dunson 1992), our risk evaluation process focused on the herpetofauna might well yield characterizations of risks that exceed expectations anticipated from tools currently applied to birds and mammals.

5.10.1 Research Needs: The Next 10 Years and Beyond Our understanding of the biology of amphibians and reptiles has continued to increase in the past dozen years, perhaps outpacing the development of our capabilities to analyze exposure and effects of environmental chemicals when herpetofauna and stressors cross paths in the field. For example, research focused on the reproductive physiology and endocrinology of amphibians and reptiles has continued to develop over the past 2 decades, which has proven beneficial to recent ecotoxicological studies focused on endocrine disruptors and herpetofauna exposed in laboratory and field. Yet, additional work is required to better characterize endpoints and mechanisms of actions of endocrine disruptors relative to monitoring activities intended to benefit adaptive management programs across a wide range of field applications, such as discharges of treated wastewaters. Beyond traditional, survival-based studies that often dominate screening level evaluations of risks, alternative, yet complementary endpoints must be refined or developed anew, particularly given the increasing awareness that life history attributes of the herpetofauna may require scrutiny and caution when comparing exposure and effects for environmental chemicals across a range of animal classes. For example, developmental effects in the herpetofauna, including traditional endpoints related to growth, may afford the most critical and sensitive endpoints linked to exposure to chemical stressors in their preferred habitats. As such, biomarkers of exposure and effects linked to developmental endpoints should be developed, or at least more fully characterized, for amphibians and reptiles, and these biomarkers must then be linked with population level effects. Long-term studies that evaluate effects over multiple life stages are required. Multigenerational studies must be completed, which is a shared research need across many animal classes. Similarly, outside of the ecotoxicological application, population level studies are available for only a handful of species, with much of that work a derivative of biodiversity concerns that have become increasingly confirmed for amphibians and reptiles, since the original statements that warned of their declining populations. Although long overlooked and consistently undervalued, amphibians and reptiles have continued to gain appreciation among technical and lay communities as critical components within many aquatic, wetland, and terrestrial ecosystems. That heightened awareness among resource managers and members of the research community, however, must be matched by increased efforts to address data gaps in our existing knowledge of chemical toxicity to the wide range of species in these vertebrate groups. More importantly, the interrelationships of these animals with other ecosystem attributes and other physical and biological stessors must be characterized to enable amphibians and

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reptiles to better serve as indicators of habitat quality and ecosystems at risk. Indeed, for aquatic habitats such as wetlands, indigenous herptofauna are more important to evaluating system sustainability than presently appreciated. These animals must be more thoroughly considered in future research, particularly as that relates to enhancing our understanding of their ecotoxicology, and the role that long-term, low-level chemical exposures play in their future. At present, adopting the perspective of a physiological ecologist in conducting ecotoxicological research remains secondary to the commonly encountered application of ecotoxicology to the ecological risk assessment process. Despite the increased focus on herpetofauna in that process, data sources remain relatively scarce for the wide range of species potentially at risk to chemical exposures. And, given the ecological risk assessment tradition developed over the past 20 to 25 years, herpetofauna will likely benefit more from being considered as critical receptors in the risk assessment process, despite having their risks to chemical exposure bound by great uncertainties. In part, these uncertainties may be better addressed if the physiological ecologist weighs in on the research needed to improve the risk assessment process for herpetofauna. It is not simply a matter of collecting more threshold concentrations across a wider range of species, but we must delve into the ecological fabric that constitutes exposure. Although a quantitative energetics basis is long from being available to risk assessors, recognizing and developing tools that ensure a truly ecological basis for evaluating risks, especially within the context of multiple stressor exposures, should be fostered and developed to move exposure models beyond the simple “you are what you eat” tools commonly applied in today’s oftentimes regulatory-driven risk assessment process. Much has been accomplished since publication of Sparling et al. (2000a), but our knowledge of the ecotoxicology of the herpetofauna still lags behind that of birds and mammals. Playing “catch up,” however, should enable our developing tools and compiling research findings to better serve these long undervalued vertebrates.

Dedication We dedicate this chapter to Wes Birge, whose early work with amphibians encouraged those of us who followed. The herpetofauna have lost an advocate, and we have lost a colleague and friend.

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Webb GJW, Cooper-Preston H. 1989. Effects of incubation temperature on crocodiles and the evolution of reptilian oviparity. Am Zool 29:953–971. Weekes HC. 1935. A review of placentation among reptiles, with particular regard to the function and evolution of the placenta. Proc Zool Soc Lond 3–4:625–645. Whiteman HH. 1994. Evolution of facultative paedomorphosis in salamanders. Quart Rev Biol 69:205–221. Whiteman HH, Howard RD. 1997. Conserving alternative amphibians phenotypes: is there anybody out there? In: Lannoo MJ, editor, Status and conservation of Midwestern amphibians. Iowa City (IA): University of Iowa Press, p 318–324. Whiteman HH, Wissinger SA, Brown WS. 1996. Growth and foraging consequences of facultative paedomorphosis in the tiger salamander, Ambystoma tigrinum nebulosum. Evol Ecol 10:433–446. Wibbels T, Bull JJ, Crews D. 1994. Temperature-dependent sex determination: a mechanistic approach. J Exp Zool 270:71–78. Wibbels T, Crews D. 1992. Specificity of steroid hormone-induced sex determination in a turtle. J Endocrinol 133:121–129 Wibbels T, Crews D. 1994. Purative aromatase inhibitor induces male sex determination in a female unisexual lizard and in a turtle with temperature-dependent sex determination. J Endocrinol 141:295–299. Widdows J, Donkin P. 1991. Role of physiological energetics in ecotoxicology. Comp Biochem Physiol 100C:69–75. Wieser W, editor. 1973. Effects of temperature on ectothermic organisms. New York: Springer-Verlag. Wilczynski W. 1992. The nervous system. In: Feder ME, Burggren WW, editors, Environmental physiology of the amphibians. Chicago: University of Chicago, p 9–39. Wingfield JC, Kenagy GJ. 1986. Natural regulation of reproductive cycles. In: Pang KT, Schreibman MP, editors, Vertebrate endocrinology: Fundamentals and biomedical implications. Vol. 4, Part B: Reproduction. San Diego (CA): Academic Press, p 181–241. Winokur RM. 1973. Adaptive modifications of buccal mucosae in turtles. Am Zool 13:1347–1348. Wirz MV, Saldiva PH, Freire-Maia DV. 2005. Micronucleus test for monitoring genotoxicity of polluted river water in Rana catesbeiana tadpoles. Bull Environ Contamin Toxicol 75:1220–1227. Wright ML, Pikula A, Babski AM, Labieniec KE, Wolan RB. 1997. Effect of melatonin on the response of the thyroid to thyrotropin stimulation in vitro. Gen Comp Endocrinol 108:298–305. Xenopoulos MA, Schindler DW. 2001. Physical factors determining ultraviolet radiation flux into ecosystems. In: Cockell C, Blaustein A, editors, Ecosystems, evolution, and ultraviolet radiation. New York: Springer, p 36–62. Yang FX, Xu Y, Wen S. 2005. Endocrine disrupting effects of nonylphenol, bisphenol A, and p,p’-DDE on Rana nigromaculata tadpoles. Bull Environ Contamin Toxicol 75:1168–1175. Yaron Z. 1985. Reptilian placentation and gestation: structure, function and endocrine control. In: Gans C, Billett F, editors, Biology of the reptilia. Vol 15. New York: Wiley, p 527–603. Zhukova TI. 1987. Change in hematological indices of lacustrine frog in connection with its inhabitation of water bodies polluted by pesticides. Ekologiya 2:54–59. Zug GR, Vitt LJ, Caldwell JP. 2001. Herpetology: an introductory biology of amphibians and reptiles. New York: Academic Press.

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6

Effects of Current-Use Pesticides on Amphibians Christine M. Lehman and Bethany K. Williams

Contents 6.1 6.2 6.3 6.4

History of Studies Involving Amphibians and Pesticides..................................................... 167 Role of Pesticides in Amphibian Population Declines.......................................................... 167 Goals for This Chapter.......................................................................................................... 169 Atrazine................................................................................................................................. 188 6.4.1 Estrogenic Effects...................................................................................................... 188 6.4.2 Direct Effects............................................................................................................. 188 6.4.3 Indirect Effects.......................................................................................................... 189 6.5 Carbaryl................................................................................................................................. 190 6.6 Glyphosate............................................................................................................................. 190 6.7 Malathion............................................................................................................................... 191 6.8 Metolachlor............................................................................................................................ 192 6.9 Types of Studies Used to Examine Pesticide Effects on Amphibians................................... 193 6.10 Conclusions............................................................................................................................ 194 References....................................................................................................................................... 194

6.1 History of Studies Involving Amphibians and Pesticides For many years, amphibians were understudied in the ecotoxicological literature. In 1989, the Canadian Wildlife Service published a comprehensive review of studies examining the effects of contaminants on amphibians (Power et al. 1989). Just 10 years later, the same organization published an updated review that included twice the number of studies (Pauli et al. 2000), indicating rapid growth in the field of amphibian ecotoxicology. However, Sparling et al. (2000) point out that the number of amphibian ecotoxicological studies remains modest relative to research utilizing other taxa. Relyea and Hoverman (2006) also report that amphibian data appear to be lagging behind other taxa, despite an increasing number of ecotoxicological studies involving freshwater ecosystems in general.

6.2 Role of Pesticides in Amphibian Population Declines Populations of amphibians have been declining worldwide for a number of years (Stuart et al. 2004), and pesticides have long been suspected as being at least partially responsible (Cowman and Mazanti 2000). The effects of pesticides on nontarget organisms such as amphibians can outlast the presence of the actual chemical in the environment. “New-generation,” or current use, pesticides have largely replaced the organophosphates and chlorinated hydrocarbons that were heavily applied in the past. Although these newer products are formulated to break down quickly and be 167

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effective at lower application rates, they can persist in the environment at concentrations adequate to impact amphibians either directly or indirectly. Thus, current use pesticides remain a threat to nontarget organisms such as amphibians. Despite the improvements in formulations, pesticides remain among the most frequently detected contaminants in surface water and groundwater worldwide (Gilliom 2007). Thus, the issue of pesticide contamination is still a clearly relevant topic for amphibian ecotoxicologists. While increasing numbers of studies are focusing on pesticide effects on amphibians, few definitive links exist between pesticide contamination and actual population declines. Linking pesticide usage to amphibian declines can be problematic for a number of reasons. Boone et al. (2009) point out the following: 1) chemicals are less acutely toxic than in previous generations, and so their effects on nontarget wildlife will be more subtle; 2) species can differ with respect to their sensitivity to chemicals; 3) pesticide concentrations in the environment can fluctuate temporally and spatially; and 4) other stressors in the environment may cause declines or interact with the pesticides in unpredictable ways. Furthermore, there is a noticeable lack of long-term transgenerational studies that would shed light on the effects that larval exposure can have on adult traits (but see Rohr and Palmer 2005; Boone 2005) and subsequent population level effects. Although pesticide contamination may seem an obvious cause of declines within agricultural landscapes, several confounding factors make attributing amphibian declines directly to pesticides difficult (Bonin et al. 2007). Increased pesticide inputs in agricultural areas occur simultaneously with other amphibian stressors, such as reductions in terrestrial habitat and altered hydrology. For example, Beja and Alcazar (2003) observed that a transition from temporary to permanent bodies of water in agricultural lands was a more important indicator of amphibian population persistence than chemical contamination. Furthermore, amphibians may not predictably demonstrate negative pesticide effects in the field. Both Piha et al. (2006) and Gilliland et al. (2001) conducted surveys in Finland and the United States, respectively, and observed that amphibian malformations were similar among agricultural vs. nonagricultural areas. Additionally, Murphy et al. (2006a, 2006b, 2006c) reported no significant relationship between field concentrations of atrazine and various anuran endpoints, including testicular oocytes and plasma steroid concentrations. Specific instances of elevated pesticide concentrations within agricultural areas have, however, been linked to injury of amphibian populations. McDaniel et al. (2008) assessed amphibians from agricultural and nonagricultural areas and discovered that the number of testicular oocytes present in adult male Rana pipiens was correlated with mixtures of pesticides and nutrients, with the number of pesticides present being an important predictor. Hayes et al. (2002b, 2002c) also examined testicular abnormalities in Rana pipiens and correlated the number of testicular oocytes with high atrazine sales, as well as field-measured atrazine concentrations of greater than 0.2 ppb. Additionally, Knutson et al. (2004) reported that ponds near row crops were more turbid and had more nutrients and agricultural chemicals — all of which could reduce amphibian population sizes. While attempting to document negative effects of pesticides in agricultural areas may be intuitive, amphibian declines due to pesticide contamination may also occur in areas with little or no intensive agriculture. Millions of tons of pesticides are used each year in urban and suburban settings (Kiely et al. 2004), and the contribution of urban areas to the insecticide load of streams may be comparable to that of agricultural areas (Hoffman et al. 2000). In addition, nonagricultural amphibian habitats may be impacted by agricultural pesticides introduced by runoff, overspray, or aerial drift and deposition. Airborne pesticides can be transported great distances (Derek at al. 1990; LeNoir et al. 1999; Thurman and Cromwell 2000; Ryan and Hites 2002) and may be linked to amphibian declines, as suggested by several recent studies. Sparling et al. (2001) recorded lower cholinesterase levels in Hyla regilla collected from regions of California’s Sierra Nevada Mountains containing higher

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169

pesticide residues. Similarly, Fellers et al. (2004) observed very low Rana muscosa survival in the same region of California. Perhaps the most compelling correlations between large-scale patterns of amphibian decline and pesticide usage have been published by Davidson et al. (2001, 2002), Davidson (2004), and Davidson and Knapp (2007). These researchers used data from the Sierra Nevada Mountains to correlate recorded population declines for several amphibian species with upwind agricultural land use, even when a number of covariates, including the presence of predatory fish, were taken into consideration.

6.3 Goals for This Chapter In this chapter, we will discuss current use pesticides that are especially relevant today, from both an ecological and a practical perspective. We will focus primarily on widely used pesticides and biological endpoints directly linked to individual fitness. Despite growing recognition of the complexity of pesticide effects and our increasing sophistication in uncovering those effects, the formula for pesticide exposure remains quite simple. In order to be a legitimate concern for nontarget organisms, a pesticide must be present in an organism’s environment at levels adequate to induce a physiological response. Because patterns of use may be good indicators of environmental prevalence, we searched for amphibian studies on the most widely applied current use pesticides in the United States (arbitrarily defined as those that were applied in excess of 1,000,000 pounds active ingredient on a single crop in 2005, as reported by the USDA National Agricultural Statistics Service [2006]). Although US data were the most readily accessible in this case, the popularity of many of the pesticides holds worldwide. Given the relatively low persistence of many new-generation pesticides, it is especially important that ecotoxicological studies keep pace with changing pesticide use patterns. In the past 20 years, for example, popular herbicides such as alachlor and atrazine have been banned in the European Union, and cyanazine has been discontinued in the United States. Meanwhile, the broad-spectrum herbicide glyphosate has risen from relative obscurity to become the most commonly applied herbicide in the world (Kiely et al. 2004). Many new genetic and biochemical techniques have been applied in amphibian ecotoxicology over the past decade (e.g., DNA microarrays), strengthening our ability to detect pesticide exposure and evaluate exposure effects. Although these techniques are vital to an integrated ecotoxicology program, we chose to limit the studies discussed here to those with response variables at the level of the individual and above. We also have not attempted to duplicate coverage of endocrine-disrupting effects of pesticides, or a detailed discussion on how pesticides can interact with other factors (see Chapter 14, this volume). In general, we limited our search of the amphibian ecotoxicological literature to studies published since the last edition of this book in 2000. Using USEPA pesticide sales and usage data (Kiely et al. 2004), we chose to examine in greater detail the top 10 pesticides for which amphibian toxicological data exist. For 5 of these pesticides (the herbicides 2,4-D and acetochlor as well as 3 common fumigants), the recent literature regarding amphibians was sparse and will not be discussed in detail. In order of usage (with their rank by million pounds of active ingredient used per year in parentheses), the remaining 5 pesticides are glyphosate (1), atrazine (2), malathion (6), metolachlor-S (9), and metolachlor (10). Additionally, we elected to discuss carbaryl, one of the most widely used home and garden pesticides and the subject of considerable amphibian research (Kiely et al., 2004). While these 6 compounds may represent the most important pesticides currently being examined within the amphibian toxicological literature, many other pesticides have the potential to impact amphibian populations. Therefore, we also compiled a comprehensive summary of pesticide-related amphibian research from 2000 through 2008 (Table 6.1).

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Survival Metamorphic traits

Survival Metamorphic traits

Survival Metamorphic traits

Thyroid hormones Escape behavior

Gene expression Metamorphosis

Mortality

Metamorphic traits Gonad/thymus histology

Predator avoidance behavior

Activity level

Mortality Parasite infection rates

2,4-D

Acetochlor

Acetochlor

Acetochlor

Acrolein

Alachlor

Amitrole

Amitrole

Atrazine

Endpoint

2,4-D

Pesticide

Laboratory Chronic exposure

Laboratory

Laboratory Acute exposure

Laboratory Chronic exposure

Laboratoryn Acute exposure

Laboratory

Laboratory Acute exposure

Mesocosms

Mesocosms

Mesocosms

Test Type

na

Predator

Predator

na

na

na or T3

na

Pesticides Community structure

Community structure

Pesticides Community structure

Additional Factors

Table 6.1 Review of Amphibian Ecotoxicological Studies between 2000 and 2009

Rana clamitans

Bufo bufo

Rana temporaria

Rana pipiens

Bufo arenarum

X. laevis

Rana catesbieana

Hyla versicolor Rana pipiens

Ambystoma macuclatum Bufo americanus Hyla versicolor Pseudacris crucifer Rana pipiens R. sylvatica

Hyla versicolor Rana pipiens

Species

Tadpoles

Tadpoles

Tadpoles

Larvae to metamorphs

Tadpoles

Tadpoles through metamorphosis

Tadpoles

Tadpoles to metamorphs

Tadpoles to metamorphs

Tadpoles to metamorphs

Life Stage

Rohr et al. 2008

Mandrillon and Saglio 2007b

Mandrillon and Saglio 2007a

Hayes et al. 2006a

Venturino et al. 2007

Crump et al. 2002

Helbing et al. 2006

Relyea 2009

Relyea 2005b

Relyea 2009

Reference

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Reproduction Larval growth Adult fitness

Metamorphic traits Survival Gonadal morphology

Organogenesis

Gonadal morphology Metamorphic traits

Survival Metamorphic traits

Immune function

Susceptibility to parasitic infestation

Survival Metamorphic traits

Sodium absorption

Metamorphic traits Mortality Infection rates

Time and size at metamorphosis ATV infection Peripheral blood leukocytes

Metamorphic traits Gonad/thymus histology

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Laboratory Chronic exposure

Laboratory

Laboratory Chronic exposure

Laboratory

Mesocosm

Laboratory Chronic exposure

Laboratory Acute exposure

Mesocosms

Laboratory Chronic exposure

Laboratory Acute exposure

Laboratory Chronic exposure

Laboratory Chronic exposure

na

Nitrate Ambystoma tigrinum virus

Iridovirus infection

na

Nitrates Carbaryl

na

na

Pesticides Community structure

na

na

Predator

na

Rana pipiens

Ambystoma tigrinum

Ambystoma macrodactylum

Rana esculenta

Hyla versicolor

Rana sylvatica

Rana pipiens

Hyla versicolor Rana pipiens

Xenopus laevis

Xenopus laevis

Hyla versicolor

Xenopus laevis

Larvae to metamorphs

Larvae

Larvae

Adult

Tadpoles to metamorphs

Tadpoles

Tadpoles

Tadpoles to metamorphs

Tadpoles to adults

Tadpoles

Tadpoles to adults

Tadpoles to adults to tadpoles

(continued)

Hayes et al. 2006b

Forson and Storfer 2006b

Forson and Storfer 2006a

Cassano et al. 2006

Boone and BridgesBritton 2006

Koprivnikar et al. 2007

Brodkin et al. 2007

Relyea 2009

Oka et al. 2008

Lenkowski et al. 2008

LaFiandra et al. 2008

DuPreez et al. 2008

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Combined trematode infection

Gonadal morphology

Larval development Sexual differentiation

Survival pre- and postexposure

Cholinesterase activity

LC50

Metamorphic traits Laryngeal development Gonadal development Aromatase activity Sex steroids

Flow cytometry DNA Nuclei per cell Developmental stage

DNA characteristics Metamorphosis

Gonadal morphology

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Endpoint

Atrazine

Pesticide

Mesocosm Chronic exposure

Laboratory Chronic exposure

Laboratory Acute exposure

Laboratory Chronic exposure

Laboratory Acute exposure

Laboratory Acute exposure

Laboratory Chronic exposure

Laboratory Chronic exposure

Field correlations

Field study

Test Type

na

na

na

na

na

na

Food levels Hydroperiod

Nitrates

na

Landscape and local characteristics of sites

Additional Factors

Table 6.1 (continued) Review of Amphibian Ecotoxicological Studies between 2000 and 2009

Xenopus laevis

Bufo americanus

Xenopus laevis

Xenopus laevis

Rana catebeiana

Rana clamitans Xenopus laevis

Ambysoma barbouri

Rana pipiens

Rana clamitans R. catesbeiana R. pipiens

H. versicolor

Species

Tadpoles to metamorphs

Tadpoles to metamorphs

Various stages of tadpoles

Tadpoles to metamorphs

Tadpoles

Tadpoles

Larvae to juveniles

Tadpoles

Adult frogs

Tadpoles

Life Stage

Jooste et al. 2005

Freeman et al. 2005

Freeman and Rayburn 2005

Coady et al. 2005

Wan et al. 2006

Wacksman et al. 2006

Rohr et al. 2006

Orton et al. 2006

Murphy et al. 2006

Koprivnikar et al. 2006

Reference

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Mortality Metamorphic traits Behavior

Behavior Postmetamorphic water retention

Metamorphic traits Gonadal morphology

Survival Time and size at metamorphosis Activity and shelter use

Mortality

Development Mass Survival

Gonadal development Metamorphic traits Laryngeal morphology

Mortality Growth Activity

Mass and SVL at metamorphosis Days to metamorphosis Survival Hematocrit

Gonadal development Laryngeal size

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Laboratory Chronic exposure

Laboratory

Laboratory Chronic exposure

Laboratory Chronic exposure

Outdoor mesocosm

Laboratory Acute exposure

Laboratory

Laboratory Chronic exposure

Laboratory Chronic exposure Postmetamorph

Mesocosm Chronic exposure

na

Nitrate

Water volume Hunger

na

Carbaryl density hydroperiod

na

Food limitation drying

na

na

Competitors Predators

Xenopus laevis

Xenopus laevis

Ambystoma barbouri

Xenopus laevis

R. sphenocephala B. americanus Ambystoma maculatum A. texanum

Bufo americanus Psuedacris crucifer Rana clamitans R. sylvatica

Ambystoma barbouri

Rana clamitans

Ambystoma barbouri

Rana sylvatica

Tadpoles to metamorphs

Tadpoles

Larvae

Tadpoles to metamorphs

Tadpoles to metamorphs

Tadpoles

Embryos to metamorphs

Tadpoles to metamorph

Larvae and metamorph

Tadpoles to metamorphs

(continiued)

Hayes et al. 2002a

Sullivan and Spence 2003

Rohr et al. 2003

Carr et al. 2003

Boone and James 2003

Storrs and Kiesecker 2004

Rohr et al. 2004

Coady et al. 2004

Rohr and Palmer 2005

Rohr and Crumrine 2005

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Gonadal development Hermaphroditism

Gonadal development

Mortality Metamorphic traits Immune function Parasitic infestation

Testis development

Ovarian development

Hemoglobin Malformities Mass Mortality Swimming Performance Ventilation

Mortality Metamorphic traits Hematocrit

Hatching success Larval mortality Development time Number of metamorphs Malformation

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Atrazine

Endpoint

Atrazine

Pesticide

Laboratory then mesocosms

Laboratory Chronic exposure

Laboratory Acute

Laboratory Chronic exposure

Laboratory Chronic exposure

Field Laboratory

Laboratory Chronic exposure

Laboratory Chronic exposure Field survey

Test Type

UV MeHg chlorpyrifos

Nitrates

na

na

na

na

na

na

Additional Factors

Table 6.1 (continued) Review of Amphibian Ecotoxicological Studies between 2000 and 2009

Hyla chrysoscelis

Rana pipiens

Bufo americanus Rana pipiens R. sylvatica

Xenopus laevis

Xenopus laevis

Rana sylvatica

Rana pipiens

Rana pipiens

Species

Embryos through metamorphosis

Tadpoles to metamorphs

Embryos Larvae Adults

Tadpoles to metamophs

Tadpoles to metamophs

Tadpoles to metamorphs

Tadpoles to metamorphs

Tadpoles Adults

Life Stage

Britson and Threlkeld 2000

Allran and Karasov 2000

Allran and Karasov 2001

Tavera-Mendoza et al. 2002b

Tavera-Mendoza et al. 2002a

Kiesecker 2002

Hayes et al. 2002c

Hayes et al. 2002b

Reference

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Mortality Metamorphis traits

Survival Fertilization Swimming performance

LC50 Carboxylesterase activity

Survival Behavior Glycogen levels

Survival Metamorphic traits

Survival Metamorphic traits

Mortality Parasite infection rates

Survival Metamorphic traits

Survival Growth Skin peptide quantity

Survival Metamorphic traits

Oviposition site selection

Survival Metamorphic traits

Survival Metamorphic traits

Atrazine

Azadirachtin

Azinphosiviethyl

Basudin

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Mesocosm

Mesocosm

Mesocosm

Mesocosms

Laboratory

Mesocosms

Laboratory Chronic exposure

Mesocosms

Mesocosms

Laboratory Acute exposure

Laboratory Acute exposure

Laboratory Acute exposure

Microcosms

pH Predation

Nitrates Atrazine

na

Fertilizers Pathogen

Chytrid

Predation Competition fertilizers

na

Pesticides Community structure

na

na

na

na

na

Rana catesbieana Rana clamitans

Hyla versicolor

Hyla chrysoscelis

Rana catesbieana

Rana boylii

Ambystoma maculatum Bufo americanus Rana sphenocephala

Rana clamitans

Hyla versicolor Rana pipiens

Bufo americanus Rana clamitans

Ptychadena bibroni

Bufo viridis

Bufo marinus

Hyla versicolor

Tadpoles to metamorphs

Tadpoles to metamorphs

Adults

Tadpoles to metamorphs

Metamorphs

Tadpoles to metamorphs

Tadpoles

Tadpoles to metamorphs

Tadpoles to metamorphs

Tadpoles

Tadpoles

Tadpoles

Tadpoles to metamorphs

(continued)

Relyea 2006a

Boone and BridgesBritton 2006

Vonesh and Buck 2007

Pugis and Boone 2007

Davidson et al. 2007

Boone et al. 2007

Rohr et al. 2008a

Relyea 2009

Boone 2008

Ezemonye and Ilechie 2007

Yesilada et al. 2006

Diana et al. 2000

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Metamorph survival Metamorph growth Overwintering success

Metamorphic traits Mortality

Mortality

Metamorphic traits Lipid reserves Mortality Metamorphic success

Survival Metamorphic traits

Mortality Metamorphic traits

Survival Metamorphic traits

Mortality Metamorphic traits

Mortality Metamorphic traits

Mortality Metamorphic traits

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Endpoint

Carbaryl

Pesticide

Mesocosms

Mesocosms

Mesocosms

Mesocosm

Field study Experimental

Mesocosms

Laboratory Chronic exposure

Laboratory Acute exposure

Mesocosm Chronic exposure

Field Enclosures

Test Type

UV-B

Densit Multiple exposures

Predation

Competition Predation

Competition

Community structure

Density

na

Nitrate

Competition

Additional Factors

Table 6.1 (continued) Review of Amphibian Ecotoxicological Studies between 2000 and 2009

Rana sphenocephala

Rana clamitans

Rana catesbeiana Notophthalmus viridescens

Rana sphenocephala

Bufo woodhousii Rana sphenocephala

Ambystoma macuclatum Bufo americanus Hyla versicolor Pseudacris crucifer Rana pipiens R. sylvatica

Ambystoma maculatum

Bufo boreas

Rana clamitans

Rana blairi R. sphenocephala Bufo woodhousii

Species

Tadpoles to metamorphs

Tadpoles

Tadpoles to metamorphs

Tadpoles to metamorphs

Tadpoles to metamorphs

Tadpoles to metamorphs

Larvae to metamorph

Tadpole

Tadpole to metamorph

Juvenile

Life Stage

Bridges and Boone 2003

Boone and Bridges 2003a

Boone and Semlitsch 2003

Mills and Semlitsch 2004

Boone et al. 2004

Relyea 2005b

Metts et al. 2005

Dwyer et al. 2005

Boone et al. 2005

Boone 2005

Reference

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Mortality

Mortality Growth Activity

Survival Metamorphic response

Survival Metamorphic traits

Mortality Metamorphic traits

Survival Metamorphic response

Genetic variation in tolerance

Mortality

Survival Metamorphic traits Malformities

Genetic variation in tolerance

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Carbaryl

Laboratory

Laboratory Chronic exposure

Laboratory Acute exposure

Laboratory

Mesocosms

Mesocosms

Laboratory Field

Mesocosms

Laboratory Chronic exposure

Laboratory Acute exposure

na

na

Predation

na

Competition Multiple exposures

Competition Predation

na

Competition Pond drying

Water volume Hunger

Predators

Various Rana spp.

Rana sphenocephala

Hyla versicolor

Rana sphenocephala

Rana clamitans

Bufo woodhouseii Hyla versicolor Rana clamitans

Hyla versicolor

Bufo woodhouseii Notophthalmus viridescens Rana blairi R. clamitans R. sphenocephala

Ambystoma barbouri

Bufo americanus Hyla versicolor Rana catesbieana R. clamitans R. pipiens R. sylvatica

Tadpoles

Tadpoles to metamorphs

Tadpoles

Tadpoles

Tadpoles to metamorphs

Tadpoles to metamorphs

Tadpoles to metamorphs

Tadpoles to metamorphs

Larvae

Tadpoles

(continued)

Bridges and Semlitsch 2000

Bridges 2000

Relyea and Mills 2001

Bridges and Semlitsch 2001

Boone et al. 2001

Boone and Semlitsch 2001

Saura-Mas et al. 2002

Boone and Semlitsch 2002

Rohr et al. 2003

Relyea 2003

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Genetic basis for tolerance

Cholinesterase activity Swimming speed

Survival Metamorphic traits

96-hour median lethal concentration

Growth Time to metamorphosis

Cholinesterase activity

Cholinesterase activity Behavior Mass

Growth Swimming performance

Mortality Deformity Biochemical endpoints

Mortality

Immune response

Metamorphic traits Gonad/thymus histology

Chlorpyrifos

Chlorpyrifos

Chlorpyrifos and oxon derivatives

Chlorpyrifos

Chlorpyrifos

Chlorpyrifos

Chlorpyrifos

Chlorpyrifos

Chlorpyrifos

Cyclophosphamide

Cyfluthrin

Endpoint

Carbaryl

Pesticide

Laboratory Chronic exposure

Laboratory Chronic exposure

Laboratory

Laboratory

Laboratory Acute exposure

Laboratory Acute exposure

Laboratory Acute exposure

Laboratory

Laboratory

Mesocosms

Laboratory

Laboratory

Test Type

na

na

na

na

na

na

Atrazine

na

na

Pesticides Community structure

na

na

Additional Factors

Table 6.1 (continued) Review of Amphibian Ecotoxicological Studies between 2000 and 2009

Rana pipiens

Rana pipiens

Rana pipiens

Xenopus laevis

Xenopus laevis

Rana sphenocephala

Rana clamitans Xenopus laevis

Smilisca phaeota

Rana boylii

Hyla versicolor Rana pipiens

Acris crepitans Hyla chrysoscelis Gastrophryne olivacea Rana sphenocephala

Hyla versicolor

Species

Larvae to metamorphs

Tadpoles

Embryos

Premetamorphs and metamorphs

Tadpoles

Tadpoles

Tadpoles

Tadpoles to metamorphosis

Tadpoles

Tadpoles to metamorphs

Tadpoles

Tadpoles

Life Stage

Hayes et al. 2006a

Albert et al. 2007

Gaizick et al. 2001

Richards and Kendall 2003b

Richards and Kendall 2003a

Widder and Bidwell 2006

Wacksman et al. 2006

Gallo-Delgado et al. 2006

Sparling and Fellers 2007

Relyea 2009

Widder and Bidwell 2008

Semlitsch et al. 2000

Reference

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Hatching success Mortality Deformities Growth/development

Immune response

Immune function

96-hour LC50

Survival Metamorphic traits

Immune response

Immune function

Mortality

Survival Metamorphic traits

Mortality Metamorphic traits Behavior

Mortality Growth Activity

Survival Predator avoidance

Cypermethrin

DDT

DDT

Diazinon and oxon derivatives

Diazinon

Dieldrin

Dieldrin

Dimethoate

Endosulfan

Endosulfan

Endosulfan

Endosulfan

Laboratory Acute exposure

Laboratory Chronic exposure

Mesocosm Chronic exposure

Mesocosms

Laboratory Acute exposure

Laboratory Acute exposure

Laboratory Chronic exposure

Mesocosms

Laboratory

Laboratory Acute exposure

Laboratory Chronic exposure

Laboratory Chronic exposure

Temperature

Water volume Hunger

Competitors Predators

Pesticides Community structure

na

na

na

Pesticides Community structure

na

na

na

na

Litoria citropa

Ambystoma barbouri

Rana sylvatica

Hyla versicolor Rana pipiens

Hyla arborea

Rana pipiens

Rana pipiens

Hyla versicolor Rana pipiens

Rana boylii

Rana pipiens

Rana pipiens

Rana arvalis

Tadpoles

Larvae

Tadpoles to metamorphs

Tadpoles to metamorphs

Tadpoles

Adults

Tadpoles

Tadpoles to metamorphs

Tadpoles

Adults

Tadpoles

Eggs to metamorphs

(continued)

Broomhall 2002

Rohr et al. 2003

Rohr and Crumrine 2005

Relyea 2009

Sayim and Kaya 2006

Gilbertson et al. 2003

Albert et al. 2007

Relyea 2009

Sparling and Fellers 2007

Gilbertson et al. 2003

Albert et al. 2007

Greulich and Pflugmacher 2003

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Pheromonal system

Mortality Metamorphic traits Immune function Parasitic infestation

Activity Predator avoidance Metamorphic traits

Cholinesterase activity

Survival Metamorphic traits

Mortality Parasite infection rates

Mortality

Mortality

Mortality

Mortality Malformations

Endosulfan

Esfenvalerate

Fenpropimorph

Fenvalerate

Glyphosate

Glyphosate

Glyphosate (Vision®)

Glyphosate Glyphosate IPA Roundup® Touchdown® Roundup Bioactive

Glyphosate (Vision)

Glyphosate formulations

Endpoint

Survival Metamorphic traits Behavior

Endosulfan

Pesticide

Caged larvae in wetlands

Laboratory

Laboratory Acute exposure

Laboratory Chronic exposure

Mesocosms

Laboratory Acute exposure

Laboratory Chronic exposure

Field Laboratory

Laboratory Acute

Laboratory Acute exposure Recovery period

Test Type

na

na

na

pH Food

na

Pesticides Community structure

na

na

na

na

Competitors

Additional Factors

Table 6.1 (continued) Review of Amphibian Ecotoxicological Studies between 2000 and 2009 Species

Scinax nasicus

Rana pipiens R. clamitans

Crinia insignifera Heleioporus eyrei Limnodynastes dorsalis Litoria moorei

Rana pipiens

Rana clamitans

Hyla versicolor Rana pipiens

Haplobatrachus tigerinus

Rana temporaria

Rana sylvatica

Notophthalmus viridescens

Litoria freycineti

Life Stage

Tadpoles

Tadpoles

Tadpoles, new metamorphs Adults

Tadpoles

Tadpoles

Tadpoles to metamorphs

Tadpoles

Tadpole to metamorph

Tadpoles to metamorphs

Adults

Tadpoles

Reference

Lajmanovich et al. 2003

Thompson et al. 2004

Mann and Bidwell 1999

Chen et al. 2004

Rohr et al. forthcoming

Relyea 2009

Tilak et al. 2003

Teplitsky et al. 2005

Kiesecker 2002

Park et al. 2001

Broomhall and Shine 2003b

180 Ecotoxicology of Amphibians and Reptiles

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Mortality

Mortality

Oviposition site selection

Survival Metamorphic traits

Mortality

Survival Metamorphic traits

Survival

Survival

Mortality Metamorphic traits Tail damage Gonadal abnormalities

Glyphosate plus surfactant (Rodeo®)

Glyphosate (Roundup)

Glyphosate (Roundup)

Glyphosate

Glyphosate (Roundup)

Glyphosate (Roundup)

Glyphosate

Glyphosate

Glyphosate

Laboratory Acute (4 species) Chronic (1 species)

Laboratory Acute exposure

Mesocosm

Mesocosms

Laboratory Acute exposure

Laboratory Chronic exposure

Mesocosm

Laboratory Acute exposure

Laboratory Acute exposure

Several glyphosate formulations

na

Sediment

Community structure

Predators

na

na

na

na

Rana clamitans R. pipiens R. sylvatica B. americanus

Bufo woodhouseii Hyla versicolor Rana sylvatica

Bufo americanus Hyla versicolor Rana pipiens

Ambystoma macuclatum Bufo americanus Hyla versicolor Pseudacris crucifer Rana pipiens R. sylvatica

Rana sylvatica R. pipiens R. clamitans R. catesbeiana Bufo americanus Hyla versicolor

Rana cascadae

Hyla versicolor/ chrysoscelis

Rana sylvatica

Rana pipiens

Tadpoles and tadpoles to metamorphs

Metamorphs

Tadpoles

Tadpoles to metamorphs

Tadpoles

Tadpoles to metamorphs

Adults

Tadpoles

Tadpoles

(continued)

Howe et al. 2004

Relyea 2005c

Relyea 2005c

Relyea 2005b

Relyea 2005a

Cauble and Wagner 2005

Takahashi 2007

Comstock et al. 2007

Trumbo 2005

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Mortality Avoidance response Growth

Mortality Postexposure growth and development

Mortality

Mortality Abnormalities

96-hour median lethal concentration

Survival Metamorphic traits

Survival Metamorphic traits

Mortality Metamorphic traits

Mortality Metamorphic traits

Mortality Parasite infection rates

Mortality Growth

Glyphosate (Vision)

Glyphosate (Kleeraway®)

Glyphosate

Glyphosate (Roundup and Rodeo)

Malathion and oxon derivatives

Malathion

Malathion

Malathion

Malathion

Malathion

Malathion

Endpoint

LC50

Glyphosate (Vision)

Pesticide

Laboratory Acute exposure

Laboratory Chronic exposure

Mesocosms

Mesocosms

Mesocosms

Mesocosm

Laboratory

Laboratory Acute exposure

Laboratory Acute exposure

Laboratory Acute exposure

Field enclosures

Laboratory Acute exposure

Test Type

na

na

Predators

Density Application frequency

Pesticides Community structure

na

na

na

na

na

na

pH

Additional Factors

Table 6.1 (continued) Review of Amphibian Ecotoxicological Studies between 2000 and 2009 Species

Rana ridibunda

Rana clamitans

Bufo americanus Rana sylvatica R. pipiens

Rana sylvatica R. pipiens

Hyla versicolor Rana pipiens

Bufo americanus Rana clamitans

Rana boylii

Xenopus laevis

Litorea moorei

Pseudacris triseriata Rana blairi

Rana clamitans R. pipiens

Bufo americanus Rana clamitans R. pipiens Xenopus laevis

Life Stage

Tadpoles

Tadpoles

Tadpoles to metamorphs

Tadpoles to metamorphs

Tadpoles to metamorphs

Tadpoles to metamorphs

Tadpoles

Embryos

Tadpoles

Tadpoles

Tadpoles to metamorphs

Tadpoles

Reference

Sayim 2008

Rohr et al. 2008a

Relyea and Hoverman 2008

Relyea and Diecks 2008

Relyea 2009

Boone 2008

Sparling and Fellers 2007

Perkins et al. 2000

Giesy et al. 2000

Smith et al. 2001

Wojtaszek et al. 2004

Edginton et al. 2004

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Survival Growth Food consumption

Survival Metamorphic traits

Mortality

Behavior Cholinesterase levels

Immune function

Mortality Metamorphic traits Immune function Parasitic infestation

Equilibrium posture Mortality Growth Development

Metamorphic traits Gonad/thymus histology

Mortality Malformity Growth Reproductive measures

Malathion

Malathion

Malathion

Malathion

Malathion

Malathion

Malathion

Metalaxyl

Methoxychlor

Laboratory Acute exposure Chronic exposure

Laboratory Chronic exposure

Laboratory Chronic exposure

Field Laboratory

Laboratory Acute exposure

Field

Laboratory Acute exposure

Mesocosms

Laboratory Chronic exposure

na

na

na

na

na

na

Predation

Community structure

na

Xenopus laevis

Rana pipiens

Rana catesbeiana

Rana sylvatica

Rana pipiens

Bufo woodhousii

Bufo americanus Hyla versicolor Rana catesbieana R. clamitans R. pipiens R. sylvatica

Ambystoma macuclatum Bufo americanus Hyla versicolor Pseudacris crucifer Rana pipiens R. sylvatica

Limnonectus limnochairs

Eggs Adults

Larvae to metamorphs

Tadpoles

Tadpoles to metamorphs

Adults

Adults

Tadpoles

Tadpoles to metamorphs

Tadpoles

(continued)

Fort et al. 2004a

Hayes et al. 2006a

Fordham et al. 2001

Kiesecker 2002

Gilbertson et al. 2003

Dickerson et al. 2003

Relyea 2004b

Relyea 2005b

Gurushankara et al. 2007

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Mortality Malformity Sex ratio Gonadal measures

Hatching Startle response

Behavior

Behavior

LC50 Metamorphic traits Malformities

Survival Metamorphic traits

Metamorphic traits Gonad/thymus histology

Mortality Growth Activity

Tissue residues

Tissue residues

Tissue residues Hatching success

Tissue residue

Methoxychlor

Methoxychlor

Methoxychlor

Methyl parathion

Metolachlor

Nicosulfron

Octylphenol

Organochlorines

Organochlorines

Organochlorines

Organochlorines

Endpoint

Methoxychlor

Pesticide

Field survey

Field survey

Field survey

Field survey

Laboratory Chronic exposure

Laboratory Chronic exposure

Mesocosms

Laboratory Acute exposure Chronic exposure

Laboratory Chronic exposure

Laboratory Acute exposure

Laboratory Acute exposure

Laboratory Chronic exposure

Test Type

na

na

Inorganics

na

Water volume Hunger

na

Pesticides Community structure

na

na

Predators

na

na

Additional Factors

Table 6.1 (continued) Review of Amphibian Ecotoxicological Studies between 2000 and 2009

Rana clamitans

Ambystoma gracile Rana aurora

Rana temporaria

Rana lessonae R. esculenta

Ambystoma barbouri

Rana pipiens

Hyla versicolor Rana pipiens

Rana tigrina

Ambystoma macrodactylum

Ambystoma macrodactylum

Xenopus tropicalis

Species

Eggs Tadpoles Adults

Eggs

Tadpoles

Tadpoles Adults

Larvae

Larvae to metamorphs

Tadpoles to metamorphs

Larvae to metamorphs

Egg to tadpole

Larvae

Larvae

Egg to metamorph

Life Stage

Gillilland et al. 2001

de Solla et al. 2002

Hofer et al. 2005

Fagotti et al. 2005

Rohr et al. 2003

Hayes et al. 2006a

Relyea 2009

Kennedy and Sampath 2001

Verrell 2000

Ingermann et al. 2002

Eroschenko et al. 2002

Fort et al. 2004b

Reference

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Mortality Avoidance response Glycogen levels

Survival Metamorphic traits

Metamorphic traits Gonad/thymus histology

Metamorphic traits Gonad/thymus histology

Metamorphic traits Gonad/thymus histology

Tissue residue

Mortality Malformations

Behavior Survival Metamorphic traits

Metamorphic traits Gonad/thymus histology

Survival Metamorphic traits

Organophosphates

Permethrin

Propiconizole

S-Metalochlor

Tebupirimphos

Toxaphene

Triclopyr

Triphenyltin

l-Cyhalothrin

2,4-D Acetochlor Atrazine Carbaryl Chlorpyrifos Diazinon Endosulfan Glyphosate Malathion Metolachlor

Mesocosms

Laboratory Chronic exposure

Laboratory Chronic exposure

Laboratory Acute exposure

Field survey

Laboratory Chronic exposure

Laboratory Chronic exposure

Laboratory Chronic exposure

Mesocosm

Laboratory Acute exposure

Multiple pesticides Community structure

na

na

na

na

na

na

na

na

na

Hyla versicolor Rana pipiens

Rana pipiens

Ambystoma barbouri

Bufo americanus Rana clamitans R. pipiens Xenopus laevis

Hyla regilla

Rana pipiens

Rana pipiens

Rana pipiens

Bufo americanus Rana clamitans

Ptychadena bibroni

Tadpoles to metamorphs

Larvae to metamorphs

Larvae to metamorphs

Embryo Larvae

Tadpoles

Larvae to metamorphs

Larvae to metamorphs

Larvae to metamorphs

Tadpoles to metamorphs

Tadpoles

(continued)

Relyea 2009

Hayes et al. 2006a

Rehage et al. 2002

Edginton et al. 2003

Angermann et al. 2002

Hayes et al. 2006a

Hayes et al. 2006a

Hayes et al. 2006a

Boone 2008

Ezemonye and Ilechie 2007

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Survival Metamorphic traits

Survival Metamorphic traits

Survival Metamorphic traits

LC50

Survival Metamorphic traits

Immune function

LC50 DNA damage

Atrazine Carbaryl

Atrazine Chlorpyrifos Metolachlor

Carbaryl Malathion Permethrin

Butachlor Dichlorvos

Glyphosate Malathion

Aldicarb Atrazine Dieldrin Endosulfan Lindane Metribuzine

Imidacloprid RH-5849

Endpoint

Metamorphic traits Gonad/thymus histology

Atrazine S-Metalochlor 7 other pesticides

Pesticide

Laboratory Acute exposure

Laboratory Acute exposure

Mesocosm

Laboratory Acute exposure

Mesocosm

Laboratory Macrocosms

Mesocosm

Laboratory Chronic exposure

Test Type

na

na

Predation

na

na

na

Nitrates

na

Additional Factors

Table 6.1 (continued) Review of Amphibian Ecotoxicological Studies between 2000 and 2009 Species

Rana N. hallowell

Rana pipiens Xenopus laevis

Bufo americanus Hyla versicolor Rana pipiens

Bufo melanostictus Fejervarya multistrada Polypedates megacephalus Micohyla ornate

Bufo americanus Rana clamitans

Bufo americanus Hyla versicolor Rana catesbieana R. clamitans R. sphenocephala Psuedacris crepitans

Hyla versicolor

Rana pipiens

Life Stage

Tadpoles

Tadpoles

Tadpoles to metamorphs

Tadpoles

Tadpoles to metamorphs

Tadpoles to metamorphs

Tadpoles to metamorphs

Larvae to metamorphs

Reference

Feng et al. 2004

Christin et al. 2004

Relyea et al. 2005

Geng et al. 2005

Boone 2008

Mazanti et al. 2003

Boone and BridgesBritton 2006

Hayes et al. 2006a

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Cholinesterase activity

Immune function

Survival Growth

Tissue residues

Mortality Growth

LC50 Avoidance response Growth

Parasite load

Parasite penetration Parasite establishment and reproduction Immune function

Body weight Liver retinoid stores

Multiple pesticides

Aldicarb Atrazine Dieldrin Endosulfan Lindane Metribuzin

Carbaryl Diazinon Glyphosate Malathion

Multiple chemicals

Azoxystrobin Cyanazine Esfenvalerate MCPA Permethrin Pirimicarb

Triclopyr Butoxyethyl ester

Atrazine Metolachlor

Atrazine Metribuzin Aldicarb Endosulfan Lindane Dieldrin

Multiple pesticides

Field survey

Laboratory

Laboratory and mesocosms

Field Acute In situ enclosures

Laboratory Acute exposure Chronic exposure

Field survey

Laboratory Chronic exposure

Laboratory Chronic exposure

Field survey

na

Lungworm exposure

Trematode exposure

na

na

na

Pesticide combinations

na

na

R. catesbeiana

R. pipiens

R. sylvatica R. clamitans

Rana clamitans R. pipiens

Rana temporaria

Rana muscosa

Bufo americanus Hyla versicolor Rana catesbieana R. clamitans R. pipiens

Rana pipiens

Hyla regilla

Adults

Juveniles

Tadpoles

Larvae

Eggs to metamorphs

Adults

Tadpoles to metamorphs

Juveniles

Tadpoles

Boily et al. 2005

Gendron et al. 2003

Griggs and Belden 2008

Wojtaszek et al. 2005

Johansson et al. 2006

Fellers et al. 2004

Relyea 2004a

Christin et al. 2003

Sparling et al. 2001

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6.4 Atrazine Atrazine is one of the most widely used herbicides in the United States and worldwide. Approximately 76 million pounds of active ingredient are used on US crops annually. Atrazine is thought to have a half-life of 3 to 90 days (Solomon et al. 1996), but can persist in the environment for long periods of time at concentrations found to cause effects to amphibians (see discussion below). Research on the effects of atrazine on amphibians has grown steadily since Hayes et al. (2002a) published their findings on the effects of this pesticide on frogs at levels previously thought to be safe for wildlife. Studies involving the effects of atrazine on amphibians can be broken down into several categories: estrogenic effects, direct toxicity, and indirect toxicity. Many of these studies include more than a single factor. By far, most recent research has focused on estrogenic effects, which are covered in detail in Chapter 8 of this book.

6.4.1 Estrogenic Effects Tavera-Mendoza et al. (2002a, 2002b) were among the first to discover the estrogenic effects of atrazine on amphibian gonads. Soon thereafter, Hayes et al. (2002a) reported Xenopus laevis tadpoles exposed in the laboratory to atrazine concentrations as low as 0.01 µg/L developed reproductive anomalies, including testicular oocytes and supernumerary ovaries. Subsequent research has corroborated these findings in Xenopus, as well as in the US native Rana pipiens (Hayes et al. 2002c). Field studies were also conducted in an attempt to correlate atrazine usage in agricultural settings with incidence of reproductive abnormalities (Hayes et al. 2002b). Other investigators have failed to observe similar effects in both the field and laboratory studies (Carr et al. 2003; Coady et al. 2004; Jooste et al. 2005; Murphy et al. 2006b; Oka et al. 2008). Hayes et al. (2006b) suggest that the mechanism by which atrazine induces gonadal malformations is by decreasing levels of androgens and increasing production of estrogen. Storrs and Semlitsch (2008) suggest that sensitivity to the estrogenic effects of atrazine on amphibians may be taxon specific. They report that species exhibiting accelerated somatic development (e.g., Bufo americanus, Hyla versicolor) also have delayed ovarian development. These species demonstrate fewer effects from exposure to estrogenic compounds because sexual differentiation occurs during or after metamorphosis. They suggest that species such as Rana sphenocephala may be more susceptible to estrogenic effects because sexual differentiation occurs during the fully aquatic larval stage.

6.4.2 Direct Effects Direct effects are those defined as being toxic to the organism. Examples of direct toxicity include impairment of the immune system, changes in behavior, alterations in growth or length of the larval period, and outright mortality. Because atrazine acts on photosynthetic systems, the herbicide’s direct toxicity on amphibians is generally sublethal. In fact, atrazine has not caused direct mortality in a variety of species even at concentrations up to an order of magnitude above expected environmental concentrations (Ambystoma barbouri, Rohr et al. 2003; Bufo americanus, Allran and Karasov 2001; Hyla versicolor, Diana et al. 2000; Rana pipiens, Allran and Karasov 2000, 2001). However, Storrs and Kiesecker (2004) did observe mortality at low doses (3 ppb), and Boone and Bridges-Britton (2006) recorded an increase in mortality when atrazine and a fertilizer were combined. Atrazine can delay development of Xenopus laevis in the laboratory (Freeman and Rayburn 2005), but can increase the size at metamorphosis of Hyla versicolor (Relyea 2009) in mesocosms. It can also increase sodium absorption in adult Rana esculenta, which could lead to a disequilibrium that would increase metabolism (Cassano et al. 2006). Atrazine can also impair immune system function of amphibians. Brodkin et al. (2007) report that adult leopard frogs (Rana pipiens) exposed to environmentally realistic concentrations of atrazine exhibited increased thioglycolate-stimulated recruitment

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189

of white blood cells and decreased activity of these cells, but not outright mortality. Forson and Storfer (2006a) found that Ambystoma tigrinum virus (ATV) infection rates of A. tigrinum larvae exposed to atrazine and sodium nitrate were higher than those of control larvae or those exposed to either stressor alone. The same researchers (2006b) found similar results for A. macrodactylum when exposed to iridovirus and atrazine. In addition to increased viral infection rates, several studies have linked atrazine exposure to increased susceptibility to parasitic infection. While Griggs and Belden (2008) found that mixtures of atrazine and metolachlor did not increase parasite load in ranid tadpoles, Kiesecker (2002) found atrazine to be among the pesticides that increased trematode cyst infestation in Rana sylvatica tadpoles. These parasitic cysts have been implicated in many cases of amphibian limb deformities (Johnson et al. 2002). Koprivnikar et al. (2007) also found that Rana sylvatica tadpoles exposed to 30 µg/L atrazine had a higher number of trematode parasites (Echinostoma trivolvis) than did tadpoles exposed to 0 or 3 µg/L concentrations. However, they noted that infection rates did not differ from controls when parasites and tadpoles were exposed simultaneously to atrazine, suggesting that high atrazine concentrations may reduce infectiousness of parasites. When Rohr et al. (2008a) exposed green frog (Rana clamitans) tadpoles and parasitic cercariae to several pesticides, they found that atrazine was the only chemical that reduced cercarial survival. They suggested that when tadpoles and cercariae were exposed to atrazine simultaneously, the net effect atrazine in the environment would be increased infection rates among tadpoles. Finally, in a field study, Rohr et al. (2008b) discovered that the best predictor of trematode infection rates in Rana pipiens tadpoles was concentration of atrazine in the environment. Atrazine may also impact larval amphibian behavior. Koprivnikar et al. (2007) found that tadpole activity was not decreased by atrazine. In fact, Rohr et al. (2003) observed increases in spontaneous tadpole activity, potentially as a result of direct effects of atrazine on the nervous system. Rohr and Crumrine (2005) also found increased tadpole activity with atrazine exposure and attributed the pattern to decreased periphyton concentrations and the subsequent increase in tadpole foraging effort. The high rate of foraging noted by Rohr and Crumrine (2005) could potentially increase vulnerability to predation, if tadpoles do not decrease movement when predators are present. Allran and Karasov (2001), however, observed that Rana pipiens tadpole feeding behavior actually declined in the presence of atrazine, despite increases in buccal and thoracic ventilation. Although atrazine seldom exceeds concentrations of 30.0 µg/L in the field (Solomon et al. 1996), levels adjacent to agricultural lands have been measured at as high as 500 µg/L (Kadoum and Mock 1978) due to runoff and overspray. Many of atrazine’s effects have been observed at levels above expected environmental concentrations (EECs; e.g., Diana et al. 2000 [200 to 2000 µg/L]; Allran and Karasov 2001 [2000 µg/L]). So, while concentrations at which direct effects have been observed are possible in the field, they are not likely. The effects of atrazine exposure can extend well beyond the larval period in which most exposures occur. Rohr and Palmer (2005) found that the direct effects of atrazine can be delayed up to 8 months postexposure. In their study, Ambystoma barbouri salamanders exposed to atrazine demonstrated accelerated water loss and increased risk for desiccation 4 and 8 months postexposure. Rohr et al. (2006) subsequently found lower survival rates in Ambystoma barbouri 14 months after exposure to ≥4 ppb atrazine — a concentration just 1 ppb higher than the USEPA drinking water standard (USEPA 2006).

6.4.3 Indirect Effects Indirect effects are defined as those that affect an ecosystem component that, in turn, affects the species of interest. For example, because atrazine is an herbicide, it targets the photosynthetic systems of plants that serve as food for tadpoles. In this manner, a decrease in algal resources attributable to atrazine application may have an indirect effect on developing amphibian larvae (DeNoyelles et al. 1982).

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Boone and James (2003) found that atrazine had negative effects on the size at metamorphosis of larval Bufo americanus and Rana sphenocephala, and attributed this effect to a reduction in chlorophyll levels (a measure of periphyton concentration). They also found atrazine to lengthen the larval period for the salamander Ambystoma texanum but suggested this may be a more direct effect of atrazine toxicity. Rohr and Crumrine (2005) also showed atrazine to have an indirect effect on developing Rana sylvatica tadpoles. Tadpoles reared in mesocosms with atrazine had longer larval periods and smaller body sizes upon metamorphosis, which corresponded to a decrease in periphyton.

6.5 Carbaryl Carbaryl is an insecticide that has recently received considerable attention in the amphibian toxicological literature (Boone and Bridges 2003b). Although it is not widely used in commercial agriculture, carbaryl is one of the most widely used insecticides in the United States in the home and garden market (Kiely et al. 2004). It is relatively nontoxic to wildlife, and despite significant within- and among-species variation with respect to carbaryl sensitivity (Bridges and Semlitsch 2000, 2001), amphibians demonstrate LC50s that are 2 to 3 times higher than environmentally expected concentrations (Boone and Bridges 1999; Bridges et al. 2002; Dwyer et al. 2005). Aside from being relatively nontoxic to amphibians, carbaryl has a relatively short half-life of 1 to 4 days (Boone and Semlitsch 2002; Boone and James 2003). Therefore, concentrations of carbaryl expected in the natural environment are generally not lethal to amphibians. However, carbaryl can interact with other factors to become directly lethal. UV radiation can increase its toxicity in the laboratory (Zaga et al. 1998), although a similar increase did not occur under field conditions (Bridges and Boone 2003). In the laboratory, low, environmentally realistic concentrations of carbaryl were lethal when 6 species of tadpoles were simultaneously exposed to predators (Relyea and Mills 2001; Relyea 2003). However, when the same exposures occurred in outdoor mesocosms, this pattern was no longer evident (Relyea 2006a). The conflicting results of laboratory versus mesocosm studies emphasize the importance of examining the effects of pesticides in natural settings. The effects of carbaryl in a more natural environment (i.e., mesocosms) appear to be complex and have been studied extensively. Boone and Semlitsch (2001, 2002, 2003) and Mills and Semlitsch (2004) found that tadpole survival to metamorphosis was actually higher in ponds that had been dosed with carbaryl. Although increased survival is a counterintuitive response to pesticide exposure, this result was shown to be an indirect effect. By killing zooplankton, carbaryl reduced numbers of predators of zooplankton — many of which also feed on tadpoles. The opposite pattern was observed in newts and salamander larvae, which depend on zooplankton for a food source (Boone and James 2003; Boone and Semlitsch 2003; Boone et al. 2007). Carbaryl can also indirectly increase tadpole size at metamorphosis by removing zooplankton that would ordinarily compete with tadpoles for algal resources (e.g., Boone et al. 2007). Interactions of carbaryl with other naturally occurring stressors have also been well studied. Biotic factors such as competition (e.g., Boone and Semlitsch 2001; Mills and Semlitsch 2004; Boone et al. 2007) and predation (Relyea 2004a), and abiotic factors such as UV radiation (Bridges and Boone 2003), hydroperiod (Boone and Semlitsch 2002), disease (Davidson et al. 2007; Pugis and Boone 2007; Rohr et al. 2008b), and other contaminants (e.g., Boone and James 2003; Boone et al. 2005; Boone and Bridges-Britton 2006) can interact with carbaryl, altering the consequences of exposure for amphibians developing in a natural environment (reviewed in Boone and Bridges 2003b).

6.6 Glyphosate Glyphosate is the most widely used pesticide in the United States (Kiely et al. 2004), with sales of a global end user market value of nearly $1.5 billion in 1997. Until 2000, this chemical received little attention in the amphibian toxicological literature (but see Mann and Bidwell 1999), but has recently received considerable attention because of its dominance of the herbicide market.

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Commercial formulations generally contain substances (e.g., surfactants, carriers, corrosion inhibitors) to increase the efficacy of the active ingredients. Tests determining the toxicity of pesticides often include only the active ingredients, as these other ingredients are considered inert. However, many of these “inert” additives can possess considerable toxicity apart from the active ingredient. A growing body of work has addressed the toxicity of glyphosate end use products and associated inert ingredients. Alone, the compound glyphosate has a moderate toxicity to developing amphibian larvae (Giesy et al. 2000). However, glyphosate is never applied as the active ingredient alone, and must therefore be tested in its end use form. Glyphosate is the active ingredient of several widely used commercial formulations, including Roundup® (i.e., Vision® in Canada) and Rodeo®. Roundup products were formulated strictly for terrestrial use and include surfactants to help the herbicide adhere to vegetation, while Rodeo was formulated for aquatic environments and contains no surfactants. In a FETAX assay, Perkins et al. (2000) found that Roundup was more toxic to developing Xenopus laevis tadpoles than was Rodeo. This difference was attributed to the toxicity of the polyethoxylated tallowamine (POEA) surfactant in Roundup, which had an LC50 as low as the active ingredient itself. Rohr et al. (2008a) found no effect on survival of Rana clamitans tadpoles when using glyphosate rather than a commercially available glyphosate formulation. Howe et al. (2004) also demonstrated that end use glyphosate products appear to have greater toxicity than the active ingredient alone. In another study examining the toxicity of a glyphosate formulation, Smith (2001) found the formulation Kleeraway® to be just as toxic to Pseudacris triseriata and Rana blairi as Roundup. In addition to being acutely toxic, Roundup can increase the length of the larval period and slow the growth rate of a variety of anuran species (Howe et al. 2004; Relyea 2004b; Wojtaszek et al. 2004; Cauble and Wagner 2005) and increase susceptibility to trematode infestation in Rana clamitans (Rohr et al. 2008b). Takahashi (2007) found that female gray treefrogs (Hyla spp.) selected oviposition sites that were free from Roundup, suggesting that adults may have the ability to avoid exposure to their developing offspring. Using Roundup, Relyea (2005a) found that several species of North American tadpoles experienced mortality at concentrations that were similar to expected environmental concentrations, particularly Rana sylvatica. The effects of this formulation were even more deadly when predators were added to the system (Relyea 2005a) and at higher environmental pH values (Chen et al. 2004; Edginton et al. 2004), and were independent of the presence of a soil substrate (Relyea 2005c). There is some debate as to whether Roundup — a product formulated for application on land — can be found in aquatic habitats at concentrations that are toxic (Thompson et al. 2006; Relyea 2006b, 2006c). Unfortunately, glyphosate is not generally a part of large-scale water quality monitoring programs (Battaglin et al. 2005) and data on environmental concentrations are lacking. A few studies have documented fairly widespread presence of glyphosate and its degradates in surface water (e.g., Battaglin et al. 2005; Kolpin et al. 2006). Therefore, it is likely that many amphibian populations are exposed at least periodically to glyphosate. This may be especially true for amphibians breeding in habitats vulnerable to herbicide overspray or runoff. Thompson et al. (2004) investigated whether vegetative buffers protect amphibian habitat from overspray during applications of Vision, an end use glyphosate formulation identical to Roundup. They suggested that when Vision was sprayed according to the product label as well as Canadian environmental guidelines, harmful effects on native amphibians should be negligible.

6.7 Malathion Malathion is the most widely applied insecticide in the United States (Kiely et al. 2004). It is an acetylcholinesterase inhibitor and has a half-life of up to 26 days. Among other uses, malathion is applied to control mosquito populations, and is key in combating mosquito-borne diseases like malaria and West Nile virus.

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Malathion concentrations that are lethal to amphibians tend to be much higher than EECs (Fordham et al. 2001; Relyea, 2004a, 2004b; Gurushankara et al. 2007; Sayim 2008). Thus, most of the effects of ecologically relevant concentrations of malathion on amphibians appear to be indirect. In fact, malathion can appear to have a positive effect on amphibian survival and biomass in mesocosms, as a result of increased mortality among predaceous insects (Relyea 2005b; Relyea et al. 2005; Relyea and Hoverman 2008). However, through a series of trophic effects (e.g., mortality of zooplankton leads to increased phytoplankton levels and subsequent decreases in periphyton available for tadpole grazing), malathion can negatively impact the growth and development of Rana pipiens (Relyea and Diecks 2008; Relyea and Hoverman 2008). Gilbertson et al. (2003) found that malathion reduced immune function in Rana pipiens, but only if the frogs were exposed to pathogens prior to insecticide exposure. While recovery of the immune system was observed, it was not until 20 weeks postexposure. Dickerson et al. (2003) observed a trend toward higher cholinesterase in Bufo woodhousii caged in areas where malathion was detected. Like that of many other pesticides, malathion’s toxicity can be altered by interactions with other factors. For example, Relyea (2004b) found that malathion is twice as toxic to Hyla versicolor tadpoles when they are simultaneously exposed to predators. Boone (2008) found that malathion increased the mass at metamorphosis of Bufo americanus tadpoles, unless tadpoles were exposed simultaneously to another acetylcholinesterase inhibitor, carbaryl. When Rana clamitans tadpoles were exposed to the same combination of chemicals, mass at metamorphosis was greater than that of control animals, or when animals were exposed to either chemical alone.

6.8 Metolachlor Metolachlor is one of the most frequently detected herbicides in surface water in the Midwest (e.g., Battaglin and Goolsby 1999; Battaglin et al. 2000, 2003, 2005; Clark and Goolsby 2000), occurring in up to 100% of stream samples (Battaglin et al. 2000; Lerch and Blanchard 2003) and up to 50% of groundwater samples (Battaglin et al. 2000). Despite the widespread presence of this herbicide in aquatic habitats, very few studies have examined potential effects of exposure on amphibians. Wan et al. (2006) reported a 96-hour LC50 of 14 mg/L for Rana catesbeiana tadpoles — a value similar to those found for rainbow trout (Onchorhynchus mykush) and Chinook salmon (Oncorhynchus tshawytscha). The apparent toxicity of the end use formulation Primextra II Magnum was considerably lower (96-hour LC50 = 56 mg/L). In another study with R. catesbeiana tadpoles, the metolachlor formulation Dual-960E induced DNA damage at concentrations as low as 0.272 mg/L (Clements et al. 1997). Metolachlor can also decrease growth and cause edema in Xenopus laevis embryos, and appears to become teratogenic after degradation to 2-ethyl-6-methylaniline (Osano et al. 2002). As part of a large study involving 9 pesticides, Hayes et al. (2006a) determined that S-metolachlor (a compound consisting primarily of the more herbicidally active S-isomer pair of metolachlor) elicited no negative effects on Rana pipiens larval survival, growth, or development at a low, environmentally relevant concentration (0.1 ppb). However, S-metolachlor did increase the frequency of damage to the thymus. Interestingly, the compound also appeared to act as an “effector,” significantly enhancing the toxicity of atrazine when tadpoles were exposed to the 2 herbicides simultaneously (as either a simple mixture or the commercial atrazine-metolachlor formulation Bicep II Magnum). Mazanti et al. (2003) exposed larval anurans (Hyla versicolor) to a similar atrazine-metolachlor formulation (Bicep II) as well as the insecticide chlorpyrifos in a laboratory setting, with treatment concentrations based on runoff data. Exposure to the higher of 2 herbicide treatments (2.54 mg/L metolachlor, 2.0 mg/L atrazine) caused slower growth and modest delays in metamorphosis, while tadpoles exposed to the lower treatment level (0.25 mg/L metolachlor, 0.2 mg/L atrazine) performed similarly to control animals. All tadpoles in the high herbicide/high insecticide treatment (2.50 mg/L metolachlor, 2.0 mg/L atrazine, 1.0 mg/L chlorpyrifos) died. In a second experiment, Mazanti et al. (2003) simulated spray-overs of experimental wetlands with the same 3 active ingredients

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and then sampled naturally occurring tadpoles for 4 months postspray. Although the effects of the herbicide mixture cannot be isolated from those of the insecticide, there were pulses of mortality in both low (0.25 mg/L metolachlor, 0.2 mg/L atrazine, 0.1 mg/L chlorpyrifos) and high (2.54 mg/L metolachlor, 2.0 mg/L atrazine, 0.1 mg/L chlorpyrifos) treatments relative to controls. The above pesticides represent the chemicals that have received the greatest amount of attention in the amphibian literature. We now turn our attention to the various types of studies that have been undertaken in the amphibian ecotoxicologial literature. The types of studies using amphibians vary greatly and range from laboratory experiments to mesocosm experiments to field studies. Each type of study aims to further our understanding of the impacts of pesticides on amphibians at mechanistic and/or ecological levels.

6.9 Types of Studies Used to Examine Pesticide Effects on Amphibians Beginning with the simplest, Frog Embryo Teratogenesis Assay–Xenopus (FETAX; ASTM 1998) was developed as a standardized test to examine the effects of toxicants (e.g., pesticides, metals, effluents) on early life stage amphibian larvae. While this assay has proven a useful tool in evaluating acute effects of contaminants on survival and development, there have been several criticisms raised because the test subject (i.e., Xenopus) is in a family that is very different from most other amphibian species, which may limit the test’s relevance (Mann 2005). Other criticisms include the sensitivity of the species and the length of the assay (96 hours). AMPHITOX is a laboratory test developed by Herkovits and Perez-Coll (1999) that is similar to FETAX, with more flexibility in choice of life stage and species used during exposure (Herkovits and Perez-Coll 2003). Both FETAX and AMPHITOX are standardized tests that allow the comparison of relative toxicity of various compounds. The lack of concrete ecological relevance notwithstanding, both of these assays are useful in determining the direct effects pesticides can have on individual traits and developmental processes. Laboratory studies without formal protocols are more flexible than FETAX or AMPHITOX assays and offer a high degree of control with limited external noise, but lack the complexity of natural exposures. While laboratory studies allow us to examine the mechanisms of pesticide toxicity, there are a number of examples of laboratory results conflicting with those from more complex mesocosm studies. For example, Zaga et al. (1998) found that the toxicity of carbaryl was increased when exposed to UV light, but Boone and Bridges (2003a) found no such effect in mesocosms. Similarly, while Relyea and Mills (2001) found that the presence of predators made carbaryl more toxic in the laboratory, Boone and Semlitsch (2001, 2003) observed no such phenomenon in mesocosms. This difference emphasizes the care that must be taken when attempting to predict pesticide effects in the field using laboratory-collected data. In the study of pesticide effects, mesocosms offer a greater degree of ecological realism than laboratory studies while allowing relatively easy manipulation of multiple factors experimentally. Mesocosms are becoming an important tool in amphibian pesticide research. Boone and James (2005) note that there has been a steady increase in their use since Rowe and Dunson (1994) first drew attention to the usefulness of mesocosms in ecotoxicological testing. However, Boone and James (2005) also point out that these studies have examined only a small number of responses of a few species to a limited number of contaminants and encourage their broader use. Aquatic enclosures placed in situ can serve as a preliminary examination of whether chemical contaminants in the environment can affect amphibians under seminatural conditions (Bishop and Martinovic 2000) and have been used to demonstrate pesticide effects in nature (Kiesecker 2002; de Solla et al. 2002; Thompson et al. 2004; Wojtaszek et al. 2004; Boone et al. 2008 [golf course ponds]). However, while these studies create realistic exposures, they can suffer from high variability among experimental units.

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More recently, larger-scale field studies have become a more common way to examine pesticide effects on amphibian populations. Correlations can be made between patterns of pesticide use and observed amphibian responses (Davidson et al. 2001, 2002; Sparling et al. 2001; Hayes et al. 2002b, 2002c; Davidson 2004; Davidson and Knapp 2007). Additionally, developing amphibians can be exposed to samples taken directly from the environment (Bridges and Little 2003), or entire ponds can be manipulated in a controlled manner (Boone et al. 2004). Increasingly, the connection of studies with the natural environment is an important component of experimental design. Studies with little or no environmental relevance have limited value, regardless of how elegant the design or clear-cut the results may be. The most successful ecotoxicology programs will integrate a wide variety of methods — laboratory exposures to establish causation and mechanisms of effect, seminatural experiments to examine effects in a more realistic context, and field and landscape level studies to uncover population and community responses (Semlitsch and Bridges 2005).

6.10  Conclusions The field of amphibian ecotoxicology has undergone tremendous growth in the years since the first comprehensive review of the literature was published (Power et al. 1989). We now know more about the lethal and sublethal effects of contaminants on multiple life stages of amphibians than at any other time in history. As amphibian ecotoxicology has matured as a discipline, researchers have begun to design studies that integrate laboratory, mesocosm, and field techniques while incorporating explicit connections with the natural world. These advances will be necessary for understanding the complex effects of pesticides on amphibians. Given the reality of amphibian population declines and the widespread nature of pesticide residues, the number of pesticide studies on amphibians should continue to grow for many years. The challenges to amphibian ecotoxicology remain formidable, however. New pesticide active ingredients are constantly being developed. Even for compounds that have been in use for decades, studies frequently expose new effects and interactions, especially as we move beyond traditional endpoints of mortality, behavior, and metamorphic characteristics. The expiration of patents on popular active ingredient molecules such as glyphosate has led to a proliferation of new end use formulations, emphasizing the importance of studying the contributions of “inert” ingredients to pesticide toxicity. In addition, although studies are slowly expanding to include life stages beyond the larval period, we know almost nothing about potential transgenerational effects of pesticide exposure. Although the body of knowledge on pesticides and amphibians has exploded relative to the early days of ecotoxicology, a disconcerting number of commonly used pesticides have never been studied at all in regard to amphibians (Hayes et al. 2006a). Amphibian pesticide research must expand both the number of pesticides and the variety of species used in testing. Well-designed studies with a high degree of ecological relevance will improve our understanding of contaminant effects on nontarget organisms, while also contributing to the conservation of amphibian populations.

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Sullivan KB, Spence KM. 2003. Effects of sublethal concentrations of atrazine and nitrate on metamorphosis of the African clawed frog. Environ Toxicol Chem 22:627–635. Takahashi M. 2007. Oviposition site selection: pesticide avoidance by gray treefrogs. Environ Toxicol Chem 26:1476–1480. Tavera-Mendoza L, Ruby S, Brousseau P, Fournier M, Cyr D, Marcogliese D. 2002a. Responses of the amphibian tadpole (Xenopus laevis) to atrazine during sexual differentiation of the testis. Environ Toxicol Chem 21:527–531. Tavera-Mendoza L, Ruby S, Brousseau P, Fournier M, Cyr D, Marcogliese D. 2002b. Response of the amphibian tadpole (Xenopus laevis) to atrazine during sexual differentiation of the ovary. Environ Toxicol Chem 21:1264–1267. Teplitsky C, Piha H, Laurila A, Merila J. 2005. Common pesticide increases costs of antipredator defenses in Rana temporaria tadpoles. Environ Sci Technol 39:6079–6085. Thurman EM, Cromwell AE. 2000. Atmospheric transport, deposition, and fate of triazine herbicides and their metabolites in pristine areas at Isle Royale National Park. Environ Sci Technol 34:3079–3085. Tilak KS, Veeraiah K, Sastry LVM, Rao JV. 2003. Effect of fenvalerate technical grade on acetyl cholinesterase activity in Indian bullfrog Haplobatrachus tigerinus (Daudin). J Environ Biol 24:445–448. Thompson DG, Solomon KR, Wojtaszek BG, Edginton AN, Stephenson GR. 2006. Letter to the editor: the impact of insecticides and herbicides on the biodiversity and productivity of aquatic communities. Ecol Appl 16:2022–2027. Thompson DG, Wojtaszek BG, Staznik B, Chartrand DT, Stephenson GR. 2004. Chemical and biomonitoring to assess potential acute effects of Vision® herbicide on native amphibian larvae in forest wetlands. Environ Toxicol Chem 23:843–849. Trumbo J. 2005. An assessment of the hazard of a mixture of the herbicide Rodeo® and the non-ionic surfactant R-11® to aquatic invertebrates and larval amphibians. California Fish Game 91:38–46. [USEPA] US Environmental Protection Agency. 2006. 2006 edition of the drinking water standards and health advisories. EPA 822-R-06-013. Washington (DC): USEPA, Office of Water. Venturino A, Montagna CM, de D’Angelo AMP. 2007. Risk assessment of Magnacide® herbicide at Rio Colorado irrigation channels (Argentina). Tier 3: studies on native species. Environ Toxicol Chem 26:177–182. Verrell P. 2000. Methoxychlor increases susceptibility to predation in the salamander Ambystoma macrodactylum. Bull Environ Toxicol Chem 64:85–92. Vonesh JR, Buck JC. 2007. Pesticide alters oviposition site selection in gray treefrogs. Oecologia 154:219–226. Wacksman MN, Maul JD, Lydy MJ. 2006. Impact of atrazine and chlorpyrifos toxicity in four aquatic vertebrates. Arch Environ Contam Toxicol 51:681–689. Wan MT, Buday C, Schroeder G, Kuo J, Pasternak J. 2006. Toxicity to Daphnia magna, Hyalella azteca, Oncorhynchus kisutch, Oncorhynhcus mykiss, Oncorhynchus tshawytscha, and Rana catesbeiana of atrazine, metolachlor, simazine, and their formulated products. Bull Environ Contam Toxicol 76:52–58. Widder PD, Bidwell JR. 2006. Cholinesterase activity and behavior in chlorpyrifos-exposed Rana sphenocephala tadpoles. Environ Toxicol Chem 25:2446–2454. Widder PD, Bidwell JR. 2008. Tadpole size, cholinesterase activity, and swim speed in four frog species after sub-lethal exposure to chlorpyrifos. Aquat Toxicol 88:9–18. Wojtaszek BF, Buscarini TM, Chartrand DT, Stephenson GR, Thompson DG. 2005. Effect of Release® herbicide on mortality, avoidance response, and growth of amphibian larvae in two forest wetlands. Environ Toxicol Chem 24:2533–2544. Wojtaszek BF, Staznik B, Chartrand DT, Stephenson GR, Thompson DG. 2004. Effects of Vision® herbicide on mortality, avoidance response, and growth of amphibian larvae in two forest wetlands. Environ Toxicol Chem 23:832–842. Yesilada E, Ozmen M, Yesilada O, Mete A. 2006. Toxicity of azinphosiviethyl on Drosophila melanogaster and Bufo viridis. Fresenius Environ Bull 15:503–507. Zaga A, Little EE, Rabeni CF, Ellersieck MR. 1998. Photoenhanced toxicity of a carbamate insecticide to early life stage anuran amphibians. Environ Toxicol Chem 17:2543–2553.

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Ecotoxicology of Pesticides in Reptiles Bruce D. Pauli, Stacey Money, and Donald W. Sparling

Contents 7.1 Introduction...........................................................................................................................204 7.2 Pyrethroid Insecticides..........................................................................................................205 7.3 Organophosphorus and Carbamate Insecticides...................................................................207 7.3.1 Organophosphates.....................................................................................................207 7.3.2 Carbamates................................................................................................................209 7.3.3 Cholinesterase Inhibition Studies..............................................................................209 7.4 Piscicides............................................................................................................................... 213 7.5 Herbicides.............................................................................................................................. 213 7.6 Fungicides.............................................................................................................................. 215 7.7 Vertebrate Pest Control Agents............................................................................................. 216 7.8 Tissue Residue Data............................................................................................................... 217 7.9 Assessment and Conclusion................................................................................................... 219 Acknowledgments........................................................................................................................... 220 References....................................................................................................................................... 221 Editor’s Note:  Since the previous edition of this book was printed in 2000, there has been relatively little information published regarding the effects of nonorganochlorine (non-OC) pesticides on reptiles. The authors and editors felt that due to the paucity of recently published data on the effects of pesticides on reptiles, the completion of an entirely new chapter, which would contain sufficiently novel information to represent a new peer-reviewed publication, would not be possible. Therefore, the following chapter is a near verbatim reprint of the text of a chapter with the same title included in the first edition of the present book (Pauli and Money 2000), with 1 difference: more recent studies are inserted in the appropriate sections of the present chapter. In total there were only 23 new, open-literature publications added for this updated version of the original chapter. The chapter now reviews the literature describing the effects of non-OC pesticides on reptiles up to and including studies published through early 2010. Following the format of the original chapter, information is included on pyrethroid, organophosphate, and carbamate pesticides, and piscicides, herbicides, and fungicides, with descriptions of sublethal, lethal, and potential population level effects. It should be noted that the more recent literature has shifted in focus to examining the effects of pesticides on reptiles in the laboratory and in the field rather than simply reporting residues, as was the case prior to 2000; some of these studies are discussed elsewhere in this volume. The original chapter (Pauli and Money 2000) included extensive appendices listing pesticide residue measurements in reptiles. More than 90% of that data, however, dealt with chlorinated pesticides such as DDT, mirex, toxaphene, etc. We have not included these older studies in the present chapter.

203

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7.1 Introduction Pesticides applied to control fishes, rodents, insects, or vegetation can inadvertently harm reptiles. However, pesticides also may be used in attempts to control “nuisance” reptiles, most commonly snakes and lizards. Pesticides also are employed in reptile veterinary medicine, usually for the control of parasites. Following pesticide field applications, regardless of application procedures, exposure of reptiles generally occurs as a result of the animals’ consuming pesticide-contaminated prey or through dermal or respiratory exposure. To the year 2000 the number of reports concerning the toxicity of pesticides to reptiles was surprisingly low. In fact, there has been little research conducted on this group since an early review by Hall (1980) emphasized that more work was urgently required. Hall (1980) compiled evidence indicating that pesticides, particularly organochlorines (OCs), can kill reptiles when used in standard agricultural practices. Hall (1980) warned that the accumulated information on the effects of pesticides and other environmental contaminants on reptiles was not only severely limited, but also that the data were of questionable relevance because they had been collected using various test methods, analytical procedures, and means of reporting results. Hall (1980) concluded that “despite nearly 40 years of study, we have only scant knowledge of which chemicals may be particularly hazardous to reptiles.” He further cautioned that few generalizations can be made about reptiles from the extensive literature on the effects of contaminants on birds (but see below). In a review published a decade later, Hall and Henry (1992) noted that almost no experimental evaluations of the sensitivity of reptiles to environmental contaminants had been made. One of the objectives of the review was to assess whether reptiles would be adequately protected by the nontarget toxicity tests required at that time for the commercial registration of pesticides and other chemicals. Remarkably, there was only a single study, on 1 species of lizard, that had been conducted in a manner that allowed comparison of the sensitivity of reptiles to other vertebrate groups: Hall and Clark (1982). This single study revealed that reptiles show sensitivity similar to that of mammals and birds in terms of their response to cholinesterase-inhibiting insecticides. Hall and Henry (1992) concluded that far too little was known to safely conclude that guidelines based on tests conducted with other vertebrate taxa offer adequate protection for reptiles, and they recommended a research strategy to fill the existing knowledge gaps. A review by Lambert (1997a) was more localized in its scope: the author reviewed the published literature and other reports on the effects of pesticides on reptiles in sub-Saharan Africa. One reason given for the review was to identify gaps in the literature concerning reptiles, given that pesticide use in Africa was likely to increase substantially. The author placed special emphasis on collecting information on reptile residue burdens and on the effects of pesticides following field applications. The result was a compilation of a fragmented and somewhat anecdotal literature on the effects of pesticides on tropical reptiles. The review included reports of reptiles killed or otherwise adversely affected — either directly or indirectly through a reduction of their prey base — as a result of applications of OC insecticides such as dichlorodiphenyltrichloroethane (DDT), dieldrin, endosulfan, and toxaphene. While most of the references related to reptile mortality from OCs used to control tsetse flies (Glossina spp.) (e.g., Wilson 1972), data included from unpublished sources also revealed that organophosphorus (OP) insecticides, such as cyanophos and chlorpyrifos, and the carbamate insecticide bendiocarb, might harm reptiles following their use for insect control. The author concluded that more information was required to determine both the direct threat of pesticides to reptiles and the potential secondary poisoning of reptile predators, such as raptorial birds, from their consumption of contaminated prey. The latter concern was earlier expressed by Koeman et al. (1978). Hopkins (2000) also expressed concern about the general lack of ecotoxicological information on reptiles. He noted that life history traits, carnivorous and insectivorous food habits, relatively long lifespan, limited movements in some species, and prolonged time to maturation enhance the need for contaminant studies on reptiles as well as their potential value in monitoring contaminant exposure. Campbell and Campbell (2002) provided a brief summary of ecotoxicological studies on reptiles with an emphasis on snakes and lizards. They noted that of the 15 families of lizards, 11 had

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no contaminant data available. Snakes were not represented any better — of the 10 families in Serpentes, only 4 had contaminant studies published. The number of pesticide-related studies was a fraction of the total number of contaminants and reptiles papers published. The authors specifically cite the need for more papers on cholinesterase-inhibiting pesticides and pyrethroids. Sparling et al. (2000) and Chapter 1 of this book both point out the continued deficiency of contaminant-related, and especially pesticide-related, papers that focus on reptiles. This deficiency is made even more noteworthy due to an apparent global decline in reptile populations that is akin to that seen in amphibians (Gibbons et al. 1998). This is also in spite of the fact that there are model species of lizards that can be raised in the laboratory and are sensitive to contaminants (Talent et al. 2002), allowing ecotoxicology investigations to be completed. In summary, despite the data and information presented in the original chapter and in the reviews mentioned above, there is still a scarce amount of information available concerning the effects of contaminants, particularly modern, in-use pesticides, on reptiles. Further studies are required to determine how vulnerable these disappearing species are to contamination. The available information on specific non-OC pesticides is presented here.

7.2 Pyrethroid Insecticides Synthetic pyrethroids are neurotoxins that act on axons in the peripheral and central nervous system by interacting with sodium channels. While pyrethroids are generally used to control insects, they may also be employed in reptile veterinary medicine, and they have been studied for their ability to directly kill reptiles. As a result, there is information on the acute toxicity of pyrethroids following direct applications to the animals during reptile control programs or veterinary treatment; in addition, there is some scattered information on effects following applications for insect control. The few available studies suggest that pyrethroids can be acutely toxic when directly applied to reptiles, but field studies have not confirmed significant effects on the local reptile fauna when pyrethroids are broadcast sprayed to control insects. Abe et al. (1994) documented that the pyrethroid insecticide prallethrin (Etoc®) could kill vipers when the snakes were sprayed with an oil-based formulation. The authors concluded that pyrethroid insecticides, which show little toxicity to birds and mammals, appear to be exceedingly toxic to snakes. Further, the pyrethroid seemed to affect the nervous system of treated snakes in a manner similar to that seen with target insects. A similar conclusion was made following a study of the potential use of synthetic pyrethroids for ectoparasite control in snake and lizard veterinary medicine. Because some ticks and mites of reptiles are resistant to OP insecticides, Mutschmann (1991) decided to study the efficacy of various pyrethroids against these parasites and the ability of certain snake and lizard species to tolerate external applications of the pyrethroids. The pyrethroids examined included deltamethrin, cypermethrin, flumethrin, and permethrin. Among the snakes tested were boa constrictor (Constrictor constrictor), red-sided garter snakes (Thamnophis sirtalis parietalis), western ribbon snakes (T. proximus), rainbow boas (Epicrates cenchria), corn snakes (Elaphe guttata), and white-lipped tree vipers (Trimeresurus albolabris). Among the lizards tested were brown basilisks (Basiliscus vittatus) and leopard geckos (Eublepharis macularius). Deltamethrin, cypermethrin, and flumethrin produced toxicity in the animals at low doses (7.5 8.9 (4.7–13.2)c 82.7 (56.2–188)c 98d >100 170 2324 (1671–3234)c

Gallotia galloti Anolis caronlinensis A. carolinensis A. carolinensis G. galloti Lacerta parva A. carolinensis

Meenakshi and Karpagaganapathi 1996 Sanchez et al. 1997 Hall and Clark 1982 Hall and Clark 1982 Hall and Clark 1982 Fossi et al. 1995 Özelmas and Akay 1995 Hall and Clark 1982

Phosphamidon Parathion Methyl-parathion Trichlorfon Azinphos-methyl Malathion

3.71e 5.62e 7.89e 53.1e 79.5e 502e

Birds (14) (18) (8) (10) (6) (6)

Baril et al. 1994 Baril et al. 1994 Baril et al. 1994 Baril et al. 1994 Baril et al. 1994 Baril et al. 1994

Parathion

≈15f

Turtle Mauremys caspica

Yawetz et al. 1983

Source: From Pauli and Money (2000). a Number of species used in the calculation of the median LD50 (see Baril et al. 1994). b 24-hour value. c Determined by moving average method; values (95% confidence limits). d Estimated value; confidence limits could not be calculated. e Median LD50 calculated from various species. f Actual value was 10 mg/kg; based on the mass of turtle hard tissue composed of the carapace and plastron, the authors estimate a soft tissue LD50 of 15 mg/kg.

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Additional work on cholinesterase characterization in Gallotia galloti has been published by Sanchez-Hernandez since the publication of the first edition of this book. Sanchez-Hernandez and Sanchez (2002) found that BChE accounted for 83% of total serum cholinesterase and was essentially absent in brain tissue. Serum BChE and AChE activities increased with pH through pH 11.0, with some leveling off after pH 9.0. Brain AChE activity reached a peak at pH 9.0 and declined at higher and lower pHs. Pralidoxime (2-PAM) is known to reactivate AChE after inhibition with an orgphanophosphorus pesticide. The authors determined that BChE may also be reactivated with 2-PAM. These findings support the interpretation that BChE can be used as an effective indicator of organophosphorus pesticide toxicity. Due to its sensitivity to exposure, high activity rates, and abundance in serum, BChE may be more effective than AChE. However, BChE cannot be used to evaluate cholinesterase depression in reptilian brains. Free-ranging G. galloti were captured from reference and agricultural sites in a field study located in the Canary Islands (Sanchez-Hernandez 2003), and animals were again assayed to determine if BChE and AChE are useful for determining exposure to cholinesterase-inhibiting pesticides. Serum collected from the postorbital sinus of G. galloti was tested for reactivation with 2-PAM and for activity of the cholinesterases. Incubation of samples in warm water for an hour was used to determine whether spontaneous reactivation consistent with carbamate exposure could be determined. Mean BChE activity rates from agricultural sites were significantly lower than those from reference sites. In 1 site 4% (5/125) of the captured animals were diagnosed as being BChE inhibited because their activity levels were lower than 2 standard deviations below the control group. At another site 22% (16/73) of the animals collected from the agricultural site had depressed BChE activity. When the samples were subjected to 2-PAM reactivation, evidence for inhibition in the animals increased to 18% (9/50) and 30% (17/56) in the 2 areas, respectively. There was no evidence of BChE inhibition in the 2 reference areas. Maximum BChE inhibition observed in the study was 94%. A third study (Sanchez-Hernandez et al. 2004) demonstrated that G. galloti exposed to carbamates in the field can show spontaneous recovery of BChE. These studies confirmed that lizard BChE can be an effective diagnostic tool of OP or carbamate exposure under field conditions.

7.4 Piscicides The effects of piscicides on reptiles involve the compound rotenone. Rotenone acts by blocking reoxidation of the reduced form of nicotinamide adenine dinucleotide (NADH) by NADH-dehydrogenase, causing death through oxygen deprivation (Fontenot et al. 1994). It is often used to remove fish from managed ponds. Haque (1971) made an early observation that rotenone may be harmful to reptiles; 1 dead aquatic snake (species not given) was found in a fish-rearing pond that had been treated 48 hours earlier with approximately 1.0 mg/L of a rotenone formulation. However, another snake was seen entering the pond (to an unknown fate), and therefore the results are somewhat inconclusive. Fontenot et al. (1994) concluded that no good studies had been conducted to document the effect of rotenone applications on reptiles. They suspected, however, that animals utilizing a high degree of dermal respiration and that are slow to leave their aquatic habitat following treatment, such as certain turtles, are probably susceptible to rotenone treatments. This supposition was soon confirmed by McCoid and Bettoli (1996), who found dead and dying common mud turtles (Kinosternon subrubrum) in shallow coves around a reservoir where 3 mg/L of a 5% rotenone formulation had been applied to assess fish community structure. At least 60 turtles died in small coves (totaling only 6.7 ha surface area); the authors speculate that probably more were killed but their carcasses were not recovered.

7.5 Herbicides Prior to 2000 only 1 report was found of a study of the direct effects of herbicides on snakes. Three other reports were found that contained anecdotal information concerning possible effects on snakes, turtles, and tortoises. However, after 2000 the effects of herbicides on reptiles were the

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most active area of pesticide research for this class. Several papers were published on the effects of terrestrial herbicides, and 2 were published on aquatic herbicides. A few studies concerning the effects of the controversial herbicide atrazine and reptiles have been published recently, but since Chapter 8 of this volume reviews the effects of atrazine on amphibians and reptiles, these papers are not included here. Littrell (1983) examined the toxicity of the herbicide thiobencarb (Bolero 10G®) to garter snakes. Thiobencarb is used in rice culture, and its use in the United States may overlap the range of the giant garter snake (Thamnophis couchi gigas). To assess the risk of herbicide applications to this rare snake, the mountain garter snake (T. e. elegans) was used as a surrogate in studies of the risk of toxicity to snakes through exposure to contaminated prey in the laboratory and aerial applications in the field. In the laboratory, thiobencarb was administered in gelatin capsules implanted in the flesh of fishes fed to the snakes. Doses of 158 to 623 mg thiobencarb/kg snake body weight did not adversely affect the 5 exposed snakes. As the estimated field exposure of a typical 200-g garter snake would amount to about 1.5 mg/kg through ingestion of contaminated food, the author concluded that an adequate margin of safety exists for the snakes consuming contaminated prey following thiobencarb aerial applications. In a field study, snakes were placed in 2 traps at the edge of a field that had been treated 2 hours previously with 45 kg/ha Bolero 10G, and 2 more snakes were placed in 2 traps in a nearby unsprayed ditch. No adverse effects were noted during a 5-day exposure of the snakes to the contaminated field or during an additional 8-day observation period after the snakes had been moved back to the laboratory. Brown (1994) captured 2 adult smooth green snakes (Opheodrys vernalis) at the base of a power line pole where herbicide granules had been heavily applied to prevent growth of vines on the pole. Containers of the granular herbicide “Weed Blast-4G” and metribuzin were found nearby, and the snakes were captured among the herbicide granules, but few other details were provided. In an early study, Pierce (1958) examined the effects of 2 applications of 2,4,5-T (“Kuron”) to 2 areas along the shore of a pond. Turtles were consistently seen throughout the study period, but the results are not very conclusive; the herbicide was applied to the pond surface and may not have mixed into the pond water very thoroughly, as the applications apparently had no effect on the submersed pond weeds. However, adverse effects of the herbicides 2,4,5-T and 2,4-D on tortoises were noted by Willemsen and Hailey (1989) during a survey they conducted on the status and distribution of turtles and tortoises in Greece. While the use of paraquat and atrazine for the removal of ground vegetation had no obvious toxic effects on tortoises (the animals were observed to consume vegetation contaminated with these herbicides), almost no tortoises were subsequently seen in areas where the scrub vegetation of low terrace walls was sprayed with 2,4-D or 2,4,5-T. Although the wall vegetation provided important cover for the animals, the authors concluded that the herbicides seemed to affect the tortoises through direct toxicity rather than indirectly through a reduction in their food or cover. This conclusion was supported by frequent observations of tortoises with swollen eyes and fluid discharge from the nose. Moreover, in 1 sprayed area that was intensively studied for a 5-year period, the number of tortoises declined substantially (44% reduction in numbers), apparently through mortality. The area was mapped into sprayed and unsprayed sectors and the movement of tortoises recorded. The numbers of tortoises decreased rapidly in the sprayed areas while remaining constant in the unsprayed sectors. The authors therefore concluded that the decrease in numbers was related to direct mortality in the sprayed areas and not to migration to unsprayed areas, but they acknowledged that physical disturbance to the treated sites was also a problem. Greater detail on this population was presented by Willemsen and Hailey (2001). Jones et al. (2000) compared the effects of herbicide plus prescribed burning vs. herbicide alone on amphibian and reptile populations in an oak/hickory-dominated community in Oklahoma. They used the herbicide tebuthiuron following label application rates of 2 kg a.i./ha. Distinct differences were observed in the communities based on treatment. The control sites were mature oak/hickory

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forests, the tebuthiuron-sprayed areas were mixed shrublands dominated by red cedar (Juniperus virginiana), and the areas treated with prescribed burning and herbicide were open parklands with scattered red cedar. Reptile abundance was greatest in the control sites and least in the herbicideonly sites; however, species diversity was nearly equal in all 3 habitat types. The authors concluded that the tebuthiuron-only areas had the least habitat diversity and that tebuthiuron spraying affected the animals negatively. Glyphosate is a widely used herbicide in both terrestrial and aquatic environments. It is used extensively in agriculture as a preemergent herbicide and in ponds and lakes for controlling noxious or invasive plant species. Several studies conducted with amphibians have demonstrated that the surfactants used with glyphosate may be more toxic than the herbicide itself (see Chapter 6, this volume). In the United States, the bog turtle (Clemmys buhlenbergii) is a federally threatened species in the northeast portion of the country and is listed as endangered by several states within its distribution. It requires a habitat mosaic of open bogs, sphagnum moss, and moist grasslands. Because 1 method of maintaining these open habitats is to spray woody vegetation with herbicides, the US Fish and Wildlife Service wanted to determine if Glypro®, a formulation of glyphosate using LI700 as a surfactant, was safe to use. To test this, the red-eared slider (Trachemys scripta elegans) was used as a surrogate species (Sparling et al. 2006). Eggs were obtained from a commercial turtle farm and dipped in Glypro concentrations ranging from 1.3 to 95% Glypro and a 3% solution of LI700. These solutions resulted in calculated exposures of 0 to 11 200 mg a.i./kg of egg. The eggs were incubated at 27 °C until hatching. Turtles exposed to the highest concentration of Glypro had a significantly lower hatching success (73% compared to 100% in controls) and significantly lower body mass at hatching than controls. Genotoxicity, as determined by flow cytometry, increased with concentration of Glypro. At 6 days posthatch, turtles in the highest exposure categories had greater difficulty in righting themselves than did controls; by 9 days posthatch, there was no significant difference observed in righting ability. The authors concluded, however, that under typical spray operations there would be a low likelihood of harm from Glypro because the soil covering a turtle nest would substantially reduce exposure.

7.6 Fungicides No information on the toxicity of fungicides to reptiles following field applications was found either before or after 2000. Prior to 2000, 1 veterinarian case report on the benzimidazole fungicide thiabendazole was located, while another study dealt with physiological effects following injection of the antifungal antibiotic cyclohexamide into caimans, turtles, and chameleons. Holt et al. (1979) treated trematode infestations in 2 rat snakes (Elaphe obsoleta quadrivittata and E. o. obsoleta) with weekly doses of thiabendazole (Equizole®). Doses were as high as 110 mg/kg for 3 to 4 weeks. Although 1 of the snakes died within 24 hours of the final treatment, the authors attributed its death to the earlier infection. The other snake tolerated the thiabendazole and made a complete recovery following the treatments. The results of this case study suggest that there may be little risk to snakes as a result of their consumption of food items or contact with surfaces contaminated with this fungicide. The antifungal antibiotic cycloheximide was used in agriculture but is no longer registered for agricultural use in the United States or Canada (Tomlin 1997). Coulson and Hernandez (1971) injected 1 or 10 mg/kg cycloheximide into spectacled caimans (Caiman latirostris), red-eared sliders, and green anoles, in which it blocked protein synthesis. This resulted in increased levels of free amino acids in tissues and body fluids. In 5 caimans, a single injection of 1 mg/kg resulted in effects noticeable for weeks (blood samples were taken from the tip of the tail for 21 days). Three of the original 5 animals had died by day 21; 10 mg/kg was invariably fatal within 3 to 4 days. The fungicide methyl thiophanate (MT) is used on a variety of crops and ornamental shrubs. Studies on rats have shown that MT may have negative effects on adrenal and thyroid glands. De

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Falco et al. (2007) separated male and female lizards (Podarcis sciula) into 4 groups as controls and those exposed to 1.5 g MT applied in water to their food, cage, and substrate, with each treatment group having a 15- or 30-day exposure. At the end of the exposures, lizards were bled and euthanized. Blood was evaluated for corticosterone, ACTH, and catecholamine concentrations. Adrenal histomorphology was also examined. Corticosterone concentrations significantly increased with MT treatment compared to controls, and the animals exposed to MT for 30 days had higher corticosterone concentrations than those exposed for 15 days. ACTH showed the opposite results: ACTH levels dropped in both treatment groups compared to controls and were lower in the 30-day than in the 15-day exposure. Adrenaline concentration increased in a similar fashion as corticosterone, and noradrenalin followed the ACTH pattern. Histomorphological differences in terms of adrenaline and noradrenaline cells and steroidogenic cords were consistent with the changes in hormone concentrations. The authors concluded that MT negatively affects the normal function of adrenal glands by interfering with the balance of hormone levels, duplicating endogenous hormone function and altering adrenal histomorphology. The effects of MT on thyroid function and histology were examined by scientists within the same laboratory (Sciarrillo et al. 2008). In an acute study the authors injected lizards intraperitoneally with 350, 500, 700, 900, and 1000 mg/kg body weight. After 15 days survivors were euthanized. Mortality was observed at 500 mg/kg and increased with dose, peaking at 70% at 1000 mg/kg. The LD50 was calculated at 900 mg/kg. Other dose-dependent effects included hind limb paralysis and dyspnea.

7.7 Vertebrate Pest Control Agents Pesticide residues in various reptile tissues have been reported in the literature since the mid-1960s. Most of the data on tissue residues, however, were collected prior to 2000, when the emphasis in most reptile and contaminant studies was on determining tissue residues. Since 2000, as detailed in Chapter 1 of this volume, the emphasis of many studies has been on effects. Pauli and Money (2000) provided an extensive summary of the tissue residue data found for reptiles for both chlorinated and nonchlorinated pesticides. The vast majority of that, however, was on chlorinated pesticides. The reader is referred to Appendix 1 of that paper for more information. Campbell and Campbell (2001) reviewed residue concentrations in snakes and found only 1 paper containing information on organic contaminants that were not chlorinated. A single study was found that mentioned pyrethroid residues in a reptile following an operational application. Bennett et al. (1983) collected 1 western ribbon snake (Thamnophis p. proximus) near an Arkansas cotton field that had received an aerial application of 0.112 kg a.i./ha fenvalerate 5 days before. The fenvalerate residue in the snake was 0.12 mg/kg wet weight (skin and gastrointestinal tract removed). This level was higher than those seen in various mice, bird, and amphibian species collected at the same site. Residues in fishes were also relatively high, which may suggest that the snake was consuming contaminated fish. In a series of reports that detailed pesticide use and residues in a cotton-growing region and nearby areas of Texas, OP and OC residues were measured in lizards. Culley and Applegate (1966) reported parathion and methyl-parathion residues in tail muscle samples from 3 species of whiptail lizards (Cnemidophorus tesselatus, C. tigris, and C. inornatus) collected in cotton fields or from the adjacent desert. The lizards’ diet was mainly termites, and residues ranged from nondetectable to approximately 5.0 ppm. Eggs, however, contained up to 5 times the concentration found in the muscle tissue of gravid females. Culley and Applegate (1967) reported similar, if slightly lower, residues of the same compounds in tail, brain, liver, coelem fat, and stomach contents of the same lizard species and the same pattern of OP accumulation in the eggs. Applegate (1970) recorded 0 to 0.7 ppm methyl-parathion and 0 to 0.1 ppm parathion in 36 whole lizards of 5 different species

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trapped at Big Bend National Park, Texas. A note of caution concerning these residues was voiced by Hall (1980), however, who pointed out that most other investigators have been unable to confirm residues of these OPs in reptiles.

7.8 Tissue Residue Data A relatively large body of literature exists on the effects of vertebrate pest control agents (mainly rodenticides) on reptile populations. Some of this information concerns the application of gas fumigants, used for the control of nuisance rodents that occupy burrows, and their potential impact on burrow-dwelling reptiles. Other studies have examined pesticides for their potential to control reptiles, particularly snakes. Many of the reports, however, are anecdotal observations of the status of the local reptile fauna following the removal of rodents using pesticidelaced baits. Most of these studies have been conducted on oceanic islands that have introduced rodent or lagomorph species, and most have typically examined the response of the reptile population following removal of the mammals rather than the direct toxicity of the pesticides to the reptiles. Various non-OC pesticides have been examined for their potential as snake control chemicals. Abe et al. (1994) documented that the pyrethroid insecticide prallethrin (Etoc®) sprayed onto the vipers Agkistrodon blomhoffii brevicaudus and Trimeresurus flavoviridis in an oil-based (kerosene) spray killed these snakes, typically within 4 hours of being sprayed. At 8 hours after being sprayed, there was 100% mortality of 5 snakes of each species treated, but there was no mortality of kerosene-sprayed controls. The 2 snake species were sprayed for 1 and 4 seconds, respectively, with a 0.3% prallethrin solution discharged at approximately 50 to 80 g/second. Symptoms included tremors, hyperactivity, and repeated attempts to bite the surrounding air. On the other hand, in order to protect reptiles during vertebrate control operations, the US Fish and Wildlife Service (USFWS 1993) assessed the risk to reptiles from the application of 16 commonly used vertebrate control pesticides. A determination of risk, in terms of “no effect” or “may affect,” was based on the species’ habitat and ecology and resulting potential exposure. Three compounds, aluminum or magnesium phosphide and potassium nitrate, were determined to be potentially hazardous to several reptile species (mainly burrow-dwelling lizards, snakes, and tortoises), as it was concluded that registered use of these chemicals could constitute a threat to the continued existence of these reptile species if they were used to fumigate animal burrows or agricultural storage enclosures. The assessment of hazard was made essentially because the vulnerable animals inhabit the burrows of target species or burrows that might otherwise be fumigated during vertebrate pest control operations. Only a few laboratory studies have examined the actual toxicity to reptiles of the pesticides used in rodent control activities. Braverman (1979) administered the rodent control chemical fluoroacetamide (Compound 1081) to 2 Palestine vipers (Vipera palaestinae), a Syrian black snake (“fire racer”) (Coluber jugularis), and 2 Montpellier’s snakes (Malpolan monspessulanus). The snakes were given 1 of 4 different regimens: a single dose of 0.1 or 0.4 mg/kg (the Palestine vipers), 4 doses totaling 1.6 mg/kg (Montpellier’s snakes), or 4 doses totaling 3.2 mg/kg (the Syrian black snake). Because there was no mortality during the experiments, the author concluded that the use of Compound 1081 for rodent control in open fields was unlikely to harm snakes. This assessment was based on the fact that doses higher than those that might occur in poisoned small mammals in the field did not kill the snakes, but it did not take into account any sublethal effects. Gopher snakes (Pituophis catenifer) were fed dead or moribund mice that had consumed grainbased baits containing 9 different rodenticides (Brock 1965). Over a 2-year period, only 17 snakes were used; snakes that did not react to the administration of 1 compound were used for further tests. The author acknowledged the flaw in the study in that using the same snake in subsequent tests might present the possibility of cumulative or synergistic effects resulting from exposure to

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different toxicants. Compounds tested included sodium fluoroacetate (Compound 1080), strychnine, endrin, arsenic trioxide, zinc phosphide, thallium sulfate, and the anticoagulants Prolin, Warfarin, and Diphacin (diphacinone). There was no observable effect in snakes that consumed mice that had ingested lethal quantities of thallium sulfate or the anticoagulants. Snakes consuming the other compounds often regurgitated the mice and exhibited no further response. The rate of regurgitation was approximately 35% of the mice consumed. The 1 exception to this pattern occurred with strychnine; this compound caused tremors and irritability and the subsequent death of 5 snakes. Detailed studies have been conducted in Australia with Compound 1080 and the shingleback skink Tiliqua rugosa (McIlroy et al. 1985; Twigg et al. 1988; Twigg and Mead 1990). In 1 region of Australia, sodium fluoroacetate occurs as a secondary compound in vegetation, and the skinks, through evolutionary exposure, have developed a remarkable resistance to its toxic effects. This resistance probably arose as a means to avoid depressed fertility rather than to prevent acute intoxication (Twigg et al. 1988). Skinks collected from this region show high fluoroacetate LD50 values (500 to 800 mg of 1080/kg) compared with animals of the same species from outside the region (LD50 of approximately 200 mg/kg) or animals of a target species such as the Norway rat (Rattus norvegicus) (LD50 of 0.22 mg/kg) (Tomlin 1997). McIlroy et al. (1985) showed that other Australian reptile species had lower LD50 values than did shingleback skinks, while another skink was comparable. In their studies bearded dragons (Pogona barbata), Gould’s monitors (Varanus gouldii), and lace monitors (V. varius) all had approximate LD50s from Compound 1080 exposure of between 40 and 120 mg/kg, while the LD50 for the blotched blue-tongued lizard (Tiliqua nigrolutea) was 336 mg/kg. Despite the lower LD50 values for some species, McIlroy et al. (1985) calculated that most of the tested reptiles would have to eat large quantities of Compound 1080–impregnated bait to obtain a toxic dose, and this risk could be further reduced by decreasing the concentration of 1080 used in meat baits. Finally, Freeman et al. (1996) noted very little consumption of a cereal-based bait impregnated with Compound 1080 when the bait was offered to individuals of another skink species (Oligosoma [Leiolopisma] maaccanni) as their only food source during a 5-day laboratory study. The anticoagulant rodenticides tested in these earlier studies (e.g., warfarin and diphacinone; Brock 1965) are typical of “first-generation” anticoagulants. These compounds generally have been replaced by “second-generation” anticoagulants such as brodificoum and flocoumafen, which show increased toxicity to rodents as a result of their accumulation and persistence in the liver (Eason and Spurr 1995). These compounds also control a wider range of rodent species, including those species resistant to other anticoagulants (Tomlin 1997). Brodifacoum and flocoumafen, both coumarin anticoagulants, are used to eradicate introduced mammals on islands where the introduced species are adversely affecting endemic wildlife and plant species. In New Zealand, for instance, where there is probably the most diverse fauna of geckos and skinks of any temperate archipelago, rodenticides are commonly being used to remove human commensals, mammalian predators, and introduced herbivores to reduce pressure on the resident fauna and flora, which includes many endangered reptile species (Merton 1987; Newman 1994). Merton (1987) conducted a detailed investigation of the impacts of brodifacoum following its use to eradicate introduced rabbits from Round Island, Mauritius. The rabbits were overgrazing the island’s vegetation and affecting its exceedingly rare fauna, which includes 6 endangered reptile species, 4 of which are endemic to Round Island. While no published data on the acute toxicity of brodifacoum to reptiles existed, Merton (1987) attempted a complete eradication of rabbits from Round Island using 2 applications of Talon 20 P pelletized baits (20 ppm brodifacoum). The 2 applications occurred at 4 and 5.7 kg/ha of the bait pellets, respectively. The choice of brodifacoum was supported by the following preliminary studies: • Feeding and bait acceptance trials with both free-ranging and captive Telfair’s skinks (Leiolopisma telfairii, one of the island’s endemic reptiles) revealed little interest in brodifacoum baits on the part of the skinks.

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• The pollard/bran pelleted bait was attractive to rabbits but was observed to be of little interest to both the skinks and their insect prey (in theory minimizing the chance of secondary poisoning of the skinks through ingestion of contaminated prey). • Talon 50 P (a 50 ppm brodifacoum bait) had been used in rodent control for a decade by the New Zealand Wildlife Service with no reports of reptile mortality. • Reptiles have a distinctly different blood coagulation chemistry than do mammals, and thus should not be affected by exposure to even relatively high levels of the anticoagulant (Merton 1987). Despite all these precautions, however, skinks were observed eating rain-softened baits on Round Island, and more than 100 dead Telfair’s skinks were eventually recovered. Yet, when 10 skinks were necropsied, only one showed signs of internal hemorrhage. The author speculated that the mortality was not the result of anticoagulation; rather, the animals appeared to have had problems thermoregulating. In spite of this apparent decimation of skinks, there were no long-term impacts on the population, and a follow-up study 3 years after the rabbit eradication (North et al. 1994) recorded increased numbers of 6 reptile species, including Telfair’s skinks. An increase in reptile populations following the removal of rodents is typically the case when these second-generation anticoagulants are used (e.g., Towns 1991, 1994; Newman 1994; Eason and Spurr 1995). Thus, it can generally be concluded that the improved habitat quality on islands following the removal of rodents more than compensates for any initial negative impacts on the reptile fauna. Nevertheless, it would obviously be prudent, given Merton’s (1987) results with Telfair’s skinks, to raise a captive colony of any species whose entire population resides on an island slated for intensive anticoagulant treatment. The only post-2000 paper we were able to find on reptile control was the use of methyl bromide on brown tree snakes (Boiga irregularis; Savarie et al. 2005). The authors confined 18 snakes in cloth bags and placed them at random in a tarpaulin-covered cargo container commonly used by commercial airlines. They fumigated the container with 12 or 24 g/m3 for 1 or 2 hours. All treatments except the 12 g/m3 over 1 hour resulted in 100% mortality. The results were promising because brown tree snakes are a serious threat to the avifauna of the Pacific Islands. Except for potentially very destructive species such as the brown tree snake and highly venomous species living within human populations, the use of fumigants and poisonous baits to control reptiles is being replaced with humane removal and, if necessary, euthanasia.

7.9 Assessment and Conclusion Almost 3 decades ago, Hall (1980) summarized the available literature on the effects of pesticides on reptiles and concluded that there was very little information on the effects of environmental contaminants on reptiles. Hall (1980) also outlined research needs, including information on sublethal effects, behavioral and reproductive impacts, contaminant kinetics, the degree of cholinesterase inhibition that is diagnostic of lethal exposure to OP or carbamate insecticides, and the effects of newer compounds such as synthetic pyrethroids. The later review of Hall and Henry (1992) noted, however, that little new information had been generated, and concluded that there essentially have been no experimental assessments of reptile responeses to contaminants. Lambert (1997a) similarly reported that residue data for reptiles are lacking, effects levels are unknown, and the available information is insufficient for a complete synthesis. Additional appeals for further research were espoused by Pauli and Money (2000), Hopkins (2000) and Campbell and Campbell (2002). Despite these frequent requests, the area of pesticide ecotoxicology, indeed ecotoxicology in general, among reptiles is sorely underrepresented. Studies on the effects of pyrethroids have been published for more than 25 years and yet are few in number. Many of those that have been published have focused on 1, deltamethrin, of the 14 or so pyrethroids. More organophosphorus pesticides have been examined than pyrethroids, but most

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of the detailed studies have occurred with malathion, which is not the most toxic OP. One area that has received a fair amount of attention is the characterization of cholinesterases and their response to OP poisoning. Only 2 carbamate insecticides — carbofuran and carbaryl — have been used in tests on reptiles. While these may be the most widely used carbamates, there are others for which we have no information. We report 1 study that examined the effects of thiobencarb herbicide on lizards, while some other herbicides and some fungicides have been studied. Most studies of the individual chemicals cited above have been laboratory studies. Thus, the potential effects on reptiles of field applications of modern pesticides remain essentially unknown. Following experimental exposures in the laboratory, through consumption of pesticide-laden baits, or as a result of direct applications to the animals in veterinary medicine or during reptile control activities, OP and synthetic pyrethroid insecticides have been demonstrated to be toxic to reptiles; the sensitivity of the tested lizard and turtle species to certain OP pesticides is comparable to that of birds. But it has been difficult to demonstrate impacts following field applications of non-OC pesticides. Therefore, field studies are warranted that might help determine the potential severity of the impacts following standard applications of the newer, nonpersistent pesticides. Studies that are critical to our understanding of reptile ecotoxicology should include genotoxic effects, endocrine disruption, long-term effects, toxicity to embryos, age-related toxicity, and maternal transfer studies. Granted, many of the nonchlorinated pesticides have relatively short half-lives that may reduce the importance of chemical fate and transfer studies, but we know little about the major pathways of exposure for reptiles. For instance, what are the relative roles of dermal absorption, inhalation, or dietary exposures? The involvement of pesticides in reptile “health” issues, such as the etiology of green turtle (Chelonia mydas) fibropapillomas (Hutchinson and Simmonds 1992; Aguirre et al. 1994; Herbst and Klein 1995) or the causes of disease or mortality in tortoises (e.g., Jacobson 1994), is uncertain. There is evidence that certain groups of the more toxic vertebrate control compounds should not be used where reptiles might be present, particularly if they are applied as burrow fumigants. Rotenone applications may be harmful to turtles. Applications of bait formulations of the “secondgeneration” anticoagulants, such as brodificoum, when used to control populations of introduced rodents that are destroying reptile habitat, do more good than harm; these compounds may be toxic to reptiles that sample the baits, but the improvements in habitat quality following the removal of the rodents appear to more than offset any short-term adverse effects on the reptiles. Judicious selection of pesticide, bait type, and even bait color (Tershy and Breese 1994) may limit even the temporary, negative impacts on reptiles inhabiting an area undergoing rodent eradication. Finally, it is remarkable that no data appear to exist concerning the effects on reptiles of field applications of many groups of pesticides, including fungicides, modern herbicides (e.g., sulfonylureas), modern insecticides (e.g., microbial insecticides such as those based on Bacillus thuringiensis, viruses, or fungal agents), piscicides besides rotenone, or pesticides used as antifouling agents on boats (e.g., tributyltin) or in wood preservation (e.g., creosote). The lack of information on the risk to reptiles from field applications of modern pesticides besides rodenticides is worrisome. This lack of knowledge corresponds to our lack of understanding of reptiles in general. As Chapter 1 of this volume indicates, the status of a large number of reptile species is very poorly known.

Acknowledgments For the first edition version of this chapter we gratefully acknowledged the assistance of Julia Costain and Julie Perrault for help with data compilation and manuscript preparation, Marie Jetten for obtaining references, Jean-François Belanger for literature searches, Karin Niemann for assistance with translations, and Doug Forsyth and Steven Sheffield for helpful comments on the original draft.

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Towns DR. 1994. The role of ecological restoration in the conservation of Whitaker’s skink (Cyclodina whitakeri), a rare New Zealand lizard (Lacertilia: Scincidae). New Zealand J Zool 21:457–471. Twigg LE, King DR, Bradley AJ. 1988. The effect of sodium monofluoracetate on plasma testosterone concentration in Tiliqua rugosa (Gray). Comp Biochem Physiol 91C:343–347. Twigg LE, Mead RJ. 1990. Comparative metabolism of, and sensitivity to, fluoroacetate in geographically separated populations of Tiliqua rugosa (Gray) (Scincidae). Aust J Zool 37:617–626. [USEPA] US Environmental Protection Agency. 1986. Hazard evaluation division standard evaluation procedure. Ecological risk assessment. EPA-540/9-85-001. Washington (DC): Office of Pesticide Programs. [USFWS] US Fish and Wildlife Service. 1993. Effects of 16 vertebrate control agents on threatened and endangered species. US Fish and Wildlife Service Biological Opinion. EPA/734/R-93/900. Washington (DC): Division of Endangered Species and Habitat Conservation. Willemsen RE, Hailey A. 1989. Status and conservation of tortoises in Greece. Herpetol J 1:315–330. Willemsen RE, Hailey A. 2001. Effects of spraying the herbicides 2,4-D and 2,4,5-T on a population of tortoise Testudo hermanni in southern Greece. Environ Pollut 113:71–78. Williams D. 1989. Potential neonate mortality due to use of D-phenothrin insecticide (“aircraft spray”). Thylacinus 14:27. Wilson VJ. 1972. Observations on the effect of dieldrin on wildlife during tsetse fly Glossina morsitans control operations in eastern Zambia. Arnoldia (Rhodesia) 5:1–12. Yawetz A, Sidis I, Gasith A. 1983. Metabolism of parathion and brain cholinesterase inhibition in Aroclor 1254-treated and untreated Caspian terrapin (Mauremys caspica rivulata, Emydidae, Chelonia) in comparison with two species of wild birds. Comp Biochem Physiol 75C:377–382. Zinkl JG, Henny CJ, Shea PJ. 1979. Brain cholinesterase activities of passerine birds in forests sprayed with cholinesterase inhibiting insecticides. In: Buck WB, editor, Animals as monitors of environmental pollutants. Washington (DC): National Academy of Science, p 356–365. Zinkl JG, Roberts RB, Henny CJ, Lenhart DJ. 1980. Inhibition of brain cholinesterase activity in forest birds and squirrels exposed to aerially applied acephate. Bull Environ Contam Toxicol 24:676–668.

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8

Atrazine in the Environment and Its Implications for Amphibians and Reptiles Christine A. Bishop, Tana V. McDaniel, and Shane R. de Solla

Contents 8.1 Atrazine in the Environment................................................................................................. 227 8.1.1 Soil and Water........................................................................................................... 227 8.1.2 Biota........................................................................................................................... 230 8.2 Toxicity to Amphibians and Reptiles.................................................................................... 230 8.2.1 Amphibians: Toxicity to Eggs, Larvae, and Metamorphs......................................... 230 8.2.1.1 Acute Toxicity............................................................................................. 230 8.2.1.2 Mixtures...................................................................................................... 231 8.2.1.3 Malformations and Edema.......................................................................... 237 8.2.1.4 Parasites and Disease.................................................................................. 238 8.2.1.5 Growth and Metamorphosis....................................................................... 239 8.2.1.6 Toxicity to Adults........................................................................................240 8.2.2 Endocrine Disruption: Receptor Binding and Modes of Action............................... 241 8.2.2.1 Adrenal Function........................................................................................ 242 8.2.2.2 Sexual Development................................................................................... 242 8.2.3 Toxicity to Aquatic Communities Including Amphibians.........................................248 8.2.4 Effects of Atrazine Exposure on Reptiles................................................................. 249 8.3 Conclusions............................................................................................................................ 251 References....................................................................................................................................... 253 Atrazine is one of the most widely used pesticides on a global basis. It is the most common pesticide detected in surface and ground water in the continental United States (Gillion et al. 2007). In areas with high corn production, atrazine concentrations in streams range as high as 3- to 10-fold greater than the 3 μg/L US Environmental Protection Agency (USEPA) drinking water standard (Thurman et al. 1991; USEPA 2006). While it is not highly persistent, atrazine is used often enough and persists long enough that amphibians, and to a lesser extent reptiles, can be exposed as eggs, juveniles, and adults to intermittent yet chronic concentrations of it throughout their lifetime. The effects of atrazine on amphibians in particular have received wide attention in recent years (Kiesecker 2002; Hayes et al. 2002, 2003, 2006a, 2006b; Hayes 2004; Hecker et al. 2004, 2005a, 2005b, 2006; Rohr et al. 2008a, 2008b), to the extent that the use of atrazine has been reviewed within the United States based solely on its potential to affect gonadal development in amphibians (Steeger et al. 2007). The ecosystems and food webs inhabited by herpetofauna may also be altered by atrazine (Rohr et al. 2008a, 2008b). Here, we explore the impacts of atrazine on the life history, health, and survival of amphibians and reptiles and on the aquatic communities they inhabit.

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Table 8.1 Examples of Surface Water Concentrations (µg/L) of Atrazine

Country

Examples of Maximum Surface Water Concentrations (µg/L)

Serbia Belgium France France: Seine River England Scotland Netherlands Croatia South Africa: High velde area China: Liao-He and Yangtse Rivers China: Yange River China: Tiaozi and Zhaosutai Rivers Canada — Ontario

4.13 3.7 0.42 0.61–0.65 7.5 4.2 9.4 1.1 9.3 1.6 6.7 30–290 95

US: Michigan US: Florida Romania:   Danube River   Olt River Denmark Switzerland Rhine River:   Netherlands   Germany

250 18 1.24 4.8 7.8 0.5

Reference Gasic et al. 2002 Bintein and Devillers 1996 Bintein and Devillers 1996 Guerit et al. 2008 Bintein and Devillers 1996 Bintein and Devillers 1996 Bintein and Devillers 1996 Gojmerac et al. 2006 DuPreez et al. 2005a Gfrerer et al. 2002a Gfrerer et al. 2002b Li et al. 2007 Takacs et al. 2002 Struger et al. 2001 Murphy et al. 2006a Schuler and Rand 2008 Kaloyanova 1998

Helweg 1994 Johnen 1990 Strosser et al. 1999

340 250

The physiochemical characteristics of atrazine make it both an effective herbicide and moderately persistent in the environment. The effectiveness of attrazine and its relative affordability have led to its widespread and intensive use worldwide. Initially registered by Ciba-Geigy in 1958, the triazine herbicide atrazine was registered for use in the 1960s in the United States, and swiftly replaced 2,4-D as the dominant herbicide for use on field corn since it offered selective weed control and reduced damage to crops. By the 1970s atrazine represented over 60% of herbicide use in corn row agriculture (Takacs et al. 2002). Other chemicals in the triazine herbicide group include cyanazine and simazine. In the 1980s, the triazines were the second most commonly sold group of pesticides in the United States, with 20  0 00 tonnes of atrazine being applied to corn crops in 1985 (Giddings et al. 2005). Annual use of atrazine in North America declined in the 1990s and 2000s, but its use remains widespread and intensive globally (Table 8.1). In Europe, atrazine was used on corn, orchards, vineyards, forestry, rose cultivation, and grassland management. Its use in France was 5000 tonnes in 1986, applied to approximately 3 million ha of corn (Bintein and Devillers 1996). Concerns regarding exposure to humans and wildlife from widespread detection in American surface waters prompted a review of triazine herbicide use by the USEPA in 1994 (Solomon et al. 1996). As a result, cyanazine was phased out of use in the United States in 2002 (USEPA 1996). In the early 1990s, world consumption of atrazine was estimated at 70  000 tons/year, 90% of this applied to corn (Bintein and Devillers 1996). Atrazine has been registered for use in over 80

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countries. In Europe, atrazine use has dropped since 1989 due to regulations restricting its use. Still, in the late 1990s, the world market for atrazine amounted to over US$400 million at the user level (Hicks 1998), and worldwide use was estimated at 149  000 to 160  000 tonnes (Short and Colborn 1999). In 1997, atrazine was applied to 53% of herbicide-treated corn in the United Kingdom and was used on 109 000 ha of grassland in 1997 (European Commission 2004). However, atrazine’s use was banned throughout the European Union in 2003 (Garthwaite et al. 1997). In Asia, atrazine use is on the rise. In China, in 2002, use was estimated at 2273 tonnes per year and was expected to continue increasing (Ren et al. 2002). The widespread use of atrazine can be attributed to its persistence, affordability, minimal crop damage, and selective control of a wide variety of broadleaf and grassy weeds. Its chemical formulation is C8H14CIN5 (2-chloro-4-ethylamino-6-iso-propylanmin-s-triazine) (CAS 1912-24-9). It is sold primarily as a wettable powder or as a water-dispersible granular formulation, but is also available as a dry, flowable powder or a suspension. Where resistance has not developed, it effectively controls some broadleafed and grassy agricultural weeds such as clover, ragweed, pigweed, smart weed, wild buckwheat, lamb’s quarters mustards, and purslane. Its primary use is on corn, which accounts for over 80% of its use. However, it is also used on sugar cane, sorghum, pineapples, nursery conifers, in the forestry industry, and to control algae in ponds (Stevens and Sumner 1991). When used, atrazine is usually applied in the spring before crop emergence, dissolved in water with or without the use of oils or surfactants, and may be applied in conjunction with fertilizers. It is also occasionally used at preplanting or postemergence. Rates for application on corn typically vary between 1 and 1.6 kg ai/ha, with maximum application rates in the United States being 2.8 kg ai/ ha, although pre-1990 maximum rates were 4.48 kg ai/ha (Giddings et al. 2005). While sugar cane accounts for a small proportion of atrazine use (3% of use in the United States), the application rates for this crop are up to 11.2 kg ai/ha. As of 2008, recommended application rates of the 11 atrazine products registered for use in Canada varied between 0.5 and 1.5 kg ai/ha. The primary degradation pathway of atrazine in the environment is through microbial metabolism (Giddings et al. 2005; Takacs et al. 2002). Atrazine is chemically broken down by hydrolysis to produce hydroxyatrazine, which is not phytotoxic and therefore no longer an active herbicide. Atrazine is also metabolized by N-dealkylation via microbial action to deethylatrazine, deisopropylatrazine, and diaminochloro-s-triazine, all of which have reduced phytotoxicity (Huber 1993; Takacs et al. 2002). Atrazine has very low volatility with a vapor pressure of 0.04 mPa at 20 °C (Stevens and Sumner 1991), so loss from surface waters and soils through volatilization is minimal.

8.1 Atrazine in the Environment 8.1.1 Soil and Water In aquatic environments contaminated with atrazine, residues are primarily present in the water column and do not sorb strongly to sediments (Giddings et al. 2005). While it is relatively stable in the water column, its chemical degradation by hydrolysis may be hastened by chemical components commonly found in surface waters, such as humic and fulvic acids, which significantly reduce its half-life in the water column. It is also degraded by photolysis in surface waters. The half-life of atrazine in the water column in freshwater aquatic environments can range from 41 to 237 days, with an average of 159 days (Giddings et al. 2005). However, in a study to assess wetlands designed to mitigate atrazine concentrations, the half-life of atrazine in a constructed freshwater wetland ranged from 16 to 48 days (Moore et al. 2000). The half-life of atrazine in the water collumn is thought to be influenced by a number of factors, including temperature, light, sediment type, bacterial community, and atrazine concentrations, leading to some variation in persistence (Giddings et al. 2005).

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In estuarine systems, the half-life is reduced by increases in salinity and ranges from 3 to greater than 90 days (Solomon et al. 1996). Atrazine is persistent in soils, with degradation time (DT50) varying widely between 20 and 385 days, depending on soil composition, pH, presence of organic acids such as fulvic acid, and soil microbial activity (Huber 1993). The DT50 tends to be lower in aerobic soils, with an average of 44 days, than in anaerobic soils and sediments such as those at the bottom of ponds and wetlands, where measured DT50 averaged 228 days (Burnett et al. 2000). Atrazine primarily enters surface waters through runoff from agricultural fields (Giddings et al. 2005). Because of atrazine’s high solubility in water, it can leach from soils during heavy rains that are common in the fall and during spring runoff or during the rainy season in tropical climates. In temperate climates, rates of leaching due to runoff vary from 18% to less than 3% of the amount applied (Huber 1993). Because atrazine’s long half-life can be in excess of several months, in both soils and aquatic environments, it can persist in water between growing seasons. Other atrazine sources to aquatic habitats include wet and dry atmospheric deposition and ground water recharge. In some areas, particularly those isolated from agricultural runoff, atmospheric deposition may be the primary source of atrazine input to surface waters. Atrazine has been detected in alpine lakes in Switzerland (0.6 μg/L) and at the Experimental Lakes District in northern Ontario, Canada, both of which are isolated from agriculture. The herbicide was detected in rainwater in Norway 4 years after it was banned from use (Takacs et al. 2002). Rainwater has contained concentrations of atrazine as high as 0.45 μg/L in agricultural areas (Hall et al. 1993). Ground water recharge is another potential source of atrazine in surface waters (Hall et al. 1993). In the Thames River watershed of southwestern Ontario, Canada, concentrations as high as 95 μg/L have been measured in tributaries draining agricultural areas (Takacs et al. 2002). Farm ponds and drains that receive direct atrazine inputs from agricultural runoff may contain higher than average atrazine concentrations than other surface waters. In corn-growing areas of the midwestern United States, concentrations up to 250 μg/L have been detected in field drains and agricultural wetlands (Murphy et al. 2006a, 2006b). Up to 57 μg/L of atrazine was measured in farm ponds in southwestern Ontario (Frank et al. 1990). While the above concentrations represent maximum atrazine concentrations, recent median concentrations of atrazine are typically much lower (Eisler 1989; Giddings et al. 2005). A risk assessment with a comprehensive summary of studies in Canada and the United States reported that the median of 90th percentile concentrations for surface waters ranged from 2.46 μg/L for regions of high atrazine use and high rainfall to 0.03 μg/L for areas of low atrazine use and low rainfall. In larger water courses such as rivers and lakes, atrazine concentrations are much lower. However, they still may exceed water quality guideline concentrations established by various jurisdictions. A survey (1991 to 1993) of atrazine in surface waters in the United States indicated over 50% of 186 stream sites exceeded the USEPA maximum concentration of atrazine of 3 μg/L (Gillion et al. 2007). During a 4-year survey of tributaries of the Thames River in southwestern Ontario, Canada, atrazine concentrations exceeded the freshwater Canadian Water Quality Guideline for the protection of aquatic life of 1.8 μg/L in 5.4% of weekly samples and 49.7% of storm event samples (Takacs et al. 2002; Table 8.1). A comprehensive survey of pesticide concentrations in streams, ground water, and biota in the United States between 1992 and 2001, carried out by the US Geological Survey, indicated that atrazine was nearly ubiquitous in surface waters and ground water. Surface water sampling was conducted in 51 watersheds, at 186 stream sites, for a total of 4380 water samples, while ground water was sampled from 5047 wells across the continental United States. It was the most frequently detected pesticide in streams from agricultural areas (over 75% of samples) (from 83 agricultural sites; Gillion et al. 2007). Eighteen percent of agricultural streams had atrazine concentrations that exceeded the benchmark for acute effects on aquatic plants of 18 μg/L (USEPA 2003; Gillion et al. 2007). Six percent of agricultural streams tested exceeded the benchmark for aquatic community effects of a 60-day average atrazine concentration of 17.5 μg/L. It was detected in urban streams in over 60% of samples from

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30 urban sites. Atrazine and its metabolite deethylatrazine were also the most commonly detected pesticides in ground water, occurring in over 30% of samples (Gillion et al. 2007). The majority of information in the literature on atrazine concentrations in the environment is reported from North America; however, there are an increasing number of studies from other parts of the globe (Table 8.1). In China, measured concentrations in some of the major rivers, such as the Liao-He and Yangtse Rivers, rarely exceeded 1.6 μg/L; an exception was the Yange River, which receives inputs from pesticide manufacturing plants. In China, as elsewhere, concentrations tend to be at their highest in the late spring, during the peak time of atrazine application, and lowest in the winter, outside of the growing season (Gfrerer et al. 2002a, 2002b). Some rivers in China have become contaminated with atrazine due to accidental industrial releases. In 1997, the Tiaozi and Zhaosutai Rivers in Liaoning Province had atrazine concentrations ranging from 30 to 290 μg/L due to an accidental discharge into the rivers from an industrial leak. Atrazine concentrations in the water were so elevated that it caused widespread failure of rice crops that were irrigated with water from these rivers (Li et al. 2007). Soil concentrations of atrazine in Serbia ranged from 0.02 to 0.1 mg/kg soil within the top 15 cm of the soil (Gasic et al. 2002). In surface waters in Siberia, atrazine was at detectable concentrations in 83% of samples from agricultural areas. Atrazine was present in 60% of surface water samples from agricultural areas at 1 to 4.13 μg/L and in ground water at up to 0.3 μg/L (Gasic et al. 2002). Amphibians often live in waters where they can be exposed to concentrations of atrazine that range from below detection limit up to several hundred micrograms per liter in wetlands that are close to atrazine sources. Atrazine can persist in water between growing seasons and therefore can be present when amphibians breed in early spring, or in tropical areas at the outset of the wet season. As the agricultural growing season progresses, atrazine concentrations will fluctuate through the period of development for amphibians. While atrazine residues in surface waters are well documented within watersheds, lakes, and higher-order streams (Solomon et al. 1996), they are less well documented at sites where highest atrazine concentrations likely occur, particularly shallow, small irrigation ponds or drains within or adjacent to farms. The lack of repeated sampling throughout amphibian breeding and development periods in wetlands means that the concentration and exposure periods for amphibians and reptiles in their typical habitats remain largely unreported. This information is essential to fully assess the risk of atrazine to amphibians and reptiles in the wild. In 2003, water samples from south central Michigan were collected on a monthly basis from May to September during a study of atrazine concentrations in surface waters and biochemical response in livers of ranid frogs (Murphy et al. 2006c). Atrazine concentrations at nonagricultural sites ranged from nondetectable to 0.23 μg/L and did not exceed 1.2 μg/L at agricultural sites (Murphy et al. 2006a). In 2002, the same agricultural areas were sampled and atrazine concentrations did not exceed 2 μg/L at most sites, but a concentration of 250 μg/L was detected in 1 sample (Murphy et al. 2006b). Du Preez et al. (2005a) measured atrazine concentrations in wetlands inhabited by amphibians from the western high veld corn-growing region of South Africa during the corn growing season. Maximum atrazine concentrations ranged from 1.2 to 9.3 μg/L. A second study of atrazine concentrations in amphibian-occupied wetlands in the same corn-growing region of South Africa, outside of the growing season, found atrazine concentrations between 0.12 and 1.23 μg/L (Du Preez et al. 2005b). In Canada, Berube et al. (2005) measured atrazine concentrations in watersheds of the Yamaska River, a large river system draining agricultural areas in Southern Quebec and utilized by bullfrogs (Rana catesbeiana). Atrazine concentrations in the water ranged from below the detection limit to 220 ng/L, although they did not sample during runoff events. From 1991 to 1993, in an intensive vegetable production area of Ontario, up to 0.101 μg/L diazinon, 6.47 μg/L atrazine, and 0.210 μg/L azinphos-methyl were detected in surface waters of the Holland River during the amphibian breeding season (Bishop et al. 1999). In southern Ontario ponds that were utilized for breeding by amphibians within apple orchards, atrazine concentrations of 0.07 to 15.0 μg/L were found

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in combination with azinphos-methyl concentrations of 0.06 to 1.0 μg/L, diazinon at 0.03 to 0.78 μg/L, and endosulfan at 0.51 to 0.53 μg/L (Harris et al. 1998). These studies document the mixture of pesticides typically found in wetlands inhabited by amphibians and reptiles within agricultural watersheds. In southwestern Ontario, farm pond and drain atrazine concentrations ranged from nondetectable to 3.13 μg/L and were detected in 37 out of 40 sites, many of which were utililized for amphibian breedings (McDaniel et al 2008). In each of these studies, other pesticides, including organophosphate insecticides, other herbicides and, in some cases, endosulfan, were also detected, emphasizing the reality that amphibians are rarely exposed solely to atrazine in the environment (Harris et al. 1998; Bishop et al. 1999; Berube et al. 2005).

8.1.2 Biota Since triazines have a relatively short half-life in biological organisms, they are unlikely to bioaccumulate within food webs. Atrazine is relatively soluble in water, with a solubility of 28 mg/L at 20 °C and 70 mg/L at 25 °C. It has a moderate Kow (2.3 to 2.8), making it an unlikely candidate for bioaccumulation in the food chain. Bioconcentration factors for most organisms tested are low. In fish, these values range from 2 years

30 days

37 days plus 14 m postexposure

Exposure Time

Nuclei stage and weight *Stage of development

Clutch size F1 Testes histology (F1 and F2) Survivorship F2 Sex ratio F2

Gonad gross morphology Gonad histology Sex ratio Larynx diameter Aromatase activity *Circulating sex steriods

*Gonad histology Gonad gross morphology Sex ratio Larynx diameter

Survivorshipa Time to metamorphosisa Body sizea Hematocrita

Survivorshipa

Density-mediated effects on survivorshipa

Endpoint

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0, 10–8, 10–7, 10–6, 10–5, 10–4 M

0, 0.1, 25

0, 0.1, 0.4, 0.8, 1, 25

0, 0.1, 1, 10, 25, 200

25 0, 10, 100

0, 1, 25, 250

0, 75, 250

0, 0.1, 1, 10, 100

0, 10 In conjunction with nitrate

Goulet and Hontela 2003

Hayes et al. 2003

Hayes et al. 2006a

Hayes et al. 2002

Hayes et al. 2002 Hecker et al. 2005a

Hecker et al. 2005b

Larson et al. 1988

Oka et al. 2008

Orton et al. 2006 R. pipiens

Xenopus (wild type and ZZ all males)

A. tigrinum

Xenopus

Xenopus Xenopus

Xenopus

Xenopus

R. pipiens

Xenopus R. catesebiana

Stages 2–24

Stage 49 tadpoles to metamorphosis

Eggs to metamorphosis

Adult males

Adult males Adult males

Tadpole to metamorphosis

Stage 25 to metamorphosis

Stage 25 to metamorphosis

Fat and adrenal cell culture

10–12 weeks

36 days

46 days 50 days

50 days

1 hour

(continued)

Body size Survivorship Time to forelimb emergence *Sex ratio *Gonad histology

Aromatase activity *Sex ratio Gonad histology Vitellogenin induction (in vitro)

*Metamorphosis *Growth *Plasma corticosterone and thyroxine

Cyp 19 expression Testicular index Aromatase activit Circulating sex steriods

*Circulating sex steroid levels Testicular histology Testicular index Aromatase activity *Sex steroid levels

*Gonadal histology and morphology *Larynx diameter

*Gonad morphology and histology

*Gonad gross morphology and histology

*Corticosterone secretion *Adrenal cell viability

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0, 25

15–25

0, 200

0, 192

0, 20, 200, 2000

0, 1, 10, 25

0, 25

Tavera-Mendoza et al. 2002b

Detenbeck et al. 1996

Boone and James 2003

Britson and Threlkeld 1998

Diana et al. 2000

Jooste et al. 2005

Rohr and Crumrine 2005 R. sylvatica

Xenopus laevis

H. versicolor

R. sphenocephala B. americanus A. maculaturm A. texanum Hyla chrysoscelis

R. pipiens

Xenopus

Xenopus

Species

Stage

Free swimming to 10 months after metamorphosis Larvae

Eggs to metamorphosis

Eggs to metamorphosis

Larvae to metamorphosis

Mecocosm Studies Larvae to metamorphosis

Stage 56

Stage 56

Significant difference between endpoint in the controls versus one or more atrazine treatments.

0, 21

Tavera-Mendoza et al. 2002a

a

Dose Atrazine (µg/L)

Study

Table 8.2 (continued) Summary of Dose–Response Studies on Amphibians and Atrazine

1 year

192 days

41 days

48 hours

48 hours

Exposure Time

*Developmental stage *Response to predation Response to competition *Activity level

Metamorphosis Growth Survival Survival *Time to metamorphosis *Body size Response to competition Deformities *Time to metamorphosis Survivorship *Body size *Body size Survivorship Time to metamorphosis Gonad histology Gonad morphology

*Ovarian histology

*Testicular histology *Testicular index

Endpoint

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A comparative examination of the acute toxicity of atrazine, alachlor, and a 50:50 mixture of the 2 chemicals to early and late larval stages of Rana pipiens and Bufo americanus and to rainbow trout (Onchorynchus mykiss) and channel catfish (Ictalurus punctatus) revealed that effects of the mixture were greater than additive for most exposures, with 96-hour LC50s of 1.5 mg/L for latestage Bufo americanus larvae. Rainbow trout and channel catfish were less sensitive than amphibian larvae (Howe et al. 1998). From 2004 to 2006, in the fruit-growing area of the Okanagan Valley, British Columbia, Canada, where pesticides, water chemistry, and hatching success of the great basin spadefoot (Spea intermontana), pacific tree frog (Pseudacris regilla), western toad (Bufo boreas), and Columbia spotted frog (Rana luteiventris) were measured, it was atrazine concentrations in the water that correlated most strongly with reduced hatching success in spadefoots, but there was no correlation with tree frogs (Bishop et al. 2008). Predator proof cages containing early-stage eggs were placed in ponds in nonagricultural reference sites and in ponds in conventionally sprayed and organic orchards. Twenty pesticides were detected in the sprayed ponds. Of these, 4 were herbicides including atrazine. Pesticides in pond water occurred at parts per trillion concentrations as high as 1410 ng/L for diazinon, 25.3 ng/L atrazine, and 57 ng/L endosulfan-sulfate in sprayed sites. Chloride, sulfate, conductivity, nitrate, and phosphorus showed significant differences among sprayed, organic, and reference sites. Great basin spadefoot mean hatching success ranged from 0 to 92% among sprayed orchards, whereas the range was 48 to 98.6% among organic orchards and 51 to 95.5% among reference sites. Mean hatching success for Pacific tree frog ranged from 22.1 to 76.1% among sprayed orchards, whereas the range was 83.4 to 97.1% among reference sites. Hatching success of western toad eggs in 2004 was as low as 0.6% in sprayed orchards and as high as 96% in organic orchards. For Columbia spotted frog in 2006, mean hatching success ranged from 0 to 67.6% among sprayed and 83.8 to 95.2% among reference sites. Variables that correlated negatively with amphibian hatching success included 12 pesticides and 7 water chemistry parameters. However, for spadefoots, stepwise regression found that, in 2005, atrazine accounted for 79% of the variation in hatching success and, in 2006, atrazine, total nitrate, and chlorpyrifos accounted for 80%. For Pacific tree frogs there were no significant correlations with pesticide concentrations. Rather, hatching success correlated with water chemistry parameters (Bishop et al. 2008). These findings and others (see also Section 8.2.1.4) emphasize the need to measure all pesticides occurring in the environment as well as water chemistry parameters during field studies that attempt to quantify impacts of any single pesticide on amphibian populations. In a study examining interactions of atrazine and chlorpyrifos in 4 aquatic vertebrates, organisms were exposed to binary mixtures of these chemicals in bioassays (Wacksman et al. 2006). Atrazine alone did not affect organisms at concentrations up to 5000 μg/L; however, the presence of atrazine at 1000 μg/L did result in a significant increase in the acute toxicity of chlorpyrifos in Xenopus laevis tadpoles. Mixed results were found with Pimephales promelas, fathead minnow, with some bioassays showing greater than additive toxicity, while others showed an additive response. No effect of atrazine on chlorpyrifos toxicity was observed for bluegill (Lepomis macrochirus) or Rana clamitans tadpoles. 8.2.1.3 Malformations and Edema There are some limited data suggesting high-dose atrazine exposure may result in deformities and/ or edema in amphibians. Howe et al. (1998) reported the appearance of abdominal edema in earlystage Rana pipiens and Bufo americanus tadpoles (Gosner 29) exposed to atrazine concentrations ranging from 2800 to 23  000 μg/L for 96 hours, although the rates of edema were not included in the report. Birge et al. (1980, 1983) reported a 3% deformity rate in Rana catesbeiana tadpoles exposed to 410 μg/L atrazine. Exposure to atrazine at 20  000 μg/L resulted in significant deformity rates in Rana pipiens, Rana sylvatica, and Bufo americanus, with a no adverse effects level (NOAEL) for

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deformity estimated at 2590 μg/L and the lowest adverse effects level (LOAEL) at 4320 μg/L (Birge et al. 1980, 1983). Deformities mainly consisted of abnormal tail shapes in tadpoles, including wavy tail and lateral flexure of the tail. Exposures of streamside salamanders (Ambystoma barbouri) to 4, 40, and 400 μg/L of atrazine for 37 days, including the larval period, did not result in any increase in deformities (Rohr et al. 2003). There was a negative correlation between the number of eye deformities and atrazine concentration in a mixed exposure study (atrazine, chlorpyrifos, and monosodium methanearsonate) (Britson and Trelkeld 1998). There was also a dose-dependent increase of deformities in Rana pipiens, Rana sylvatica, and Bufo americanus larvae with increasing atrazine concentration (0.02 to 20 mg/L) (Allran and Karasov 2001). In a static renewal experimental treatment, the effects of 10, 25, and 35 mg/L atrazine from early organ morphogenesis through the onset of tadpole feeding were measured in larval Xenopus laevis (Lenkowski et al. 2008). There were significant dosedependent increases in the percentage of atrazine-exposed tadpoles with malformations of multiple tissues, including the main body axis, circulatory system, kidney, and digestive system. Incidence of apoptotic cells also increased in midbrains and kidneys of atrazine-exposed tadpoles (Lenkowski et al. 2008). 8.2.1.4 Parasites and Disease Primary hosts for trematodes are typically mollusks, while larval anurans are susceptible to trematode infections as secondary intermediate hosts. Parasitism may potentially lead to reduced survivorship and hind limb deformities. Kiesecker (2002) studied Rana sylvatica tadpoles by caging them within ponds exposed to agricultural runoff, including atrazine. An increased rate of trematode infections and hind limb deformities was found compared to tadpoles raised in ponds that did not receive agricultural inputs. Rana sylvatica tadpoles exposed to atrazine in ponds adjacent to agricultural fields developed a significant number of limb deformities. Rana sylvatica from agricultural runoff ponds exposed to cercariae were 37% smaller than their counterparts in the same ponds not exposed to cercariae And, in the agricultural runoff ponds, a significantly higher percentage (28.6%) of the Rana Sylvatica exposed to trematode infection developed limb deformities, compared with 0% among tadpoles shielded from trematode infection (Kiesecker 2002). In the same study, a dose-response exposure was also reported in which Rana sylvatica tadpoles were exposed, individually, to atrazine, malathion, and esfenvalerate, at 3 or 30 μg/L for 4 weeks, and then cercariae of 2 species of trematodes, Ribeiroia and Telorchis, were exposed at both concentrations of atrazine (Kiesecker 2002). Rana sylvatica tadpoles exhibited statistically reduced immune response, as measured by the number of circulating leukocytes, and a significant increase in trematode infection (Kiesecker 2002). At all but the lowest exposure to malathion, all 3 pesticides had a similar effect on eosinophil numbers and proportion of cercariae that encysted. Pesticide exposure also had a treatment comparable effect on the amphibian responses to the 2 different parasites (Kiesecker 2002). Both field and laboratory results suggest that there is some common physiological effect on the immune response among pesticides despite the differences in mechanisms of action among chemicals. Rana pipiens metamorphs were exposed to a mixture of atrazine, metribuxin, aldicarb, endosulfan, lindane, and dieldrin, representative of compounds and concentrations (ranging from 0.02 ng/L up to 21 μg/L) occurring in rivers of Quebec, Canada (Gendron et al. 2003), including 21 μg/L of atrazine. The frogs were exposed to an infection challenge with lungworms (Rhabdias ranae), a common frog parasite. Exposed frogs were infected with lungworms more quickly than nonexposed frogs. Although there was no significant difference in overall parasite burden, there were significantly more gravid female parasites found in the lungs of frogs from the highest-dosage group, suggesting that the exposure of the host to pesticides had altered the life history of the parasite, accelerating its life cycle. In a related experiment with Rana pipiens and Xenopus laevis, Christin et al. (2004) used the same pesticide exposure regimen as Gendron et al. (2003). There were no effects on cellularity of

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immune tissues and responses in Rana pipiens even at the highest pesticide concentrations, but the mixture significantly decreased the number of splenocytes in Xenopus. A significant reduction in the number of phagocytes with a dose-response trend was also found for Xenopus. In contrast, suppression of T-lymphocyte proliferation occurred at all concentrations in Rana pipiens and then recovered 3 weeks after exposure ended, while no modulations were observed in Xenopus (Christin et al. 2003, 2004). Rana pipiens were then challenged with Rhabdias ranae. No pesticide effects on phagocytosis and splenocyte numbers were detectable at the end of the exposure period, but these 2 parameters were diminished 21 days after the infection challenge in frogs previously exposed to the higher concentratons of pesticide mixture (Christin et al. 2003). Results at 21 μg/L atrazine, in combination with other pesticides, are further supported by studies of exposure solely to atrazine (21 μg/L for 8 days), which affected the innate immune response of adult Rana pipiens in ways similar to acid exposure (pH 5.5) (Brodkin et al. 2007). Atrazine suppressed the thioglycollate-stimulated recruitment of white blood cells to the peritoneal cavity to “background” levels and also decreased the phagocytic activity of these cells (Brodkin et al. 2007). Sometimes the interaction between parasites and pesticide exposure is in an unexpected direction, depending on the sensitivity of both the parasite and the host. In the aformentioned studies, only the amphibian host was exposed to atrazine, not the parasite. Koprivnikar et al. (2006) exposed the cercariae (infectious stage) of 4 species of digenetic trematodes to atrazine to determine if this impacted their survivorship and ability to infect anuran larvae. Exposure to 200 μg/L atrazine led to reduced survivorship and motility in some of the trematode species. There was also a significant reduction in infection rates of Rana clamitans tadpoles by exposed cercariae. This may counteract effects seen in the previous studies. When the hosts alone (Rana sylvatica) were exposed to atrazine at 3 and 30 μg/L, the rate of trematode infection was increased compared to Rana sylvatica raised without atrazine (Koprinvikar et al. 2007). When both hosts (Rana sylvatica) and trematode cercariae were exposed to 3 and 30 μg/L atrazine, infection rates in Rana sylvatica were no different from those raised without atrazine (Koprinvikar et al. 2007). Thus, it appears that atrazine may compromise both the amphibian host’s ability to mount an immune response and the ability of the parasite to infect the host. Similarly, in a short-term exposure of parasites and tadpoles to a mixture of metolachlor and atrazine, effects on the parasites and its snail host were more pronounced than in tadpoles and their susceptibility to infection. Changes in survivorship occurred in free-living trematode cercariae in low (10 and 15 μg/mL) and high (85 and 100 μg/mL) concentrations of mixtures of metolachlor and atrazine. A significant decline in cercarial survivorship in the high-concentration treatments at 14 hours was found. In a second experiment, the parasites, the second intermediate host tadpoles (Rana sylvatica and Rana clamitans), or both parasites and tadpoles were exposed to those mixtures for up to 10 hours. The mixtures had no significant effects on parasite load, but newly shed cercariae were more likely than 10-hour-old cercariae to infect tadpoles. They also used outdoor mesocosms to expose parasites, infected snail hosts, and Rana sylvatica tadpoles to those pesticide mixtures and found no significant effects on tadpole parasite loads in mesocosms (Griggs and Belden 2008). Forson and Storfer (2006) exposed 6-week-old long-toed salamander larvae (Ambystoma macrodactylum) to Ambystoma tigrinum iridovirus (ATV) in conjunction with sublethal concentrations of atrazine ranging from 1.84 to 184 μg/L. They found that infection rates of ATV and mortality from ATV were significantly lower at all 3 atrazine concentrations when larvae were exposed to ATV in conjunction with atrazine, than when larvae were exposed to ATV alone, suggesting atrazine may compromise viral efficacy. 8.2.1.5 Growth and Metamorphosis There are mixed results regarding the impact of low-dose atrazine exposure on larval growth and time to metamorphosis. These depend on the species and the concentrations of atrazine, and may

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also depend on the somatic development rate of amphibians (Ogielska and Kotusz 2004; see also Section 8.2.2.2). In short-term experiments, few significant impacts have been reported in anurans using atrazine concentrations commonly found in the environment, but toxic effects at low concentrations in a long-term study of salamanders indicate that the effects in wild amphibian populations are very complex and possibly underestimated in previous shorter-term studies (Rohr et al. 2006; Forson and Storfer 2006). Atrazine (20 to 200 μg/L), nitrate (0.5 to 20 mg/l NO3-N/L), and a mixture of these compounds had no significant effects on development rate, percent metamorphosis, time to metamorphosis, percent survival, mass at metamorphosis, or hematocrit, although nitrate slowed the growth of Rana pipiens larvae exposed at Gosner stage 25 through to metamorphosis (Allran and Karasov 2000). Similarly, exposure of juvenile Rana pipiens for 21 days to a pesticide mixture that included 21 μg/L atrazine did not result in any changes in growth or condition compared to controls (Gendron et al. 2003). There was little impact on survivorship to metamorphosis, or hind limb length as a measure of size, in caged Rana pipiens tadpoles exposed to atrazine at 15 to 75 μg/L in treated stream mesocosms, although time to metamorphosis was accelerated in atrazine-treated animals (Detenbeck et al. 1996). Larval Xenopus laevis were exposed to 1, 10, or 25 μg/L atrazine at 48 hours posthatch until metamorphosis was complete. Estradiol, dihydrotestosterone, and ethanol vehicle were also tested for effects on the larvae. None of the atrazine treatments affected posthatch mortality, larval growth, or metamorphosis (Carr et al. 2003). Similarly, larvae of Xenopus laevis were exposed to 0.01 to 200 μg/L atrazine throughout larval development. Atrazine at these concentrations had no effect on mortality, time to metamorphosis, length, or weight at metamorphosis (Hayes et al. 2002). Time to metamorphosis was not affected in exposures of Xenopus laevis to 0.01 to 100 μg/L (Hosmer et al. 2007; Steeger et al. 2007). Six-week-old Ambystoma macrodactylum larvae were exposed for 30 days to 1.84, 18.4, and 184 μg/L of atrazine (Forson and Storfer 2006). Exposure to the highest concentration of atrazine resulted in significantly decreased time to metamorphosis and a corresponding reduction in snoutvent length and body weight at metamorphosis. There was no significant impact on survivorship. When Larson et al. (1998) exposed tiger salamander (Ambystoma tigrinum) larvae to 250 μg/L atrazine, a decrease in body mass at metamorphosis was found without a change in time to metamorphosis, whereas larvae exposed to 75 μg/L metamorphosed significantly later than controls but with no reduction in body mass. Streamside salamander (Ambystoma barbouri) larvae exposed from egg stage to 37 days to 4, 40, and 400 μg/L atrazine did not show a reduction in hatching success, survivorship, or growth (Rohr et al. 2003). However, salamanders exposed to ≥4 μg/L atrazine had significantly lower survival than did control animals 14 months postexposure (Rohr et al. 2006). 8.2.1.6 Toxicity to Adults No LC50 data exist for adult amphibian exposures either orally or through skin exposures or injections. However, based on data we can extrapolate from small mammals, adverse effect doses are expected to be higher for adult animals. The oral LD50 for atrazine is 3090 mg/kg in rats, 1750 mg/ kg in mice, 750 mg/kg in rabbits, and 1000 mg/kg in hamsters (Ecotoxnet http://extoxnet.orst.edu/ pips/atrazine.htm). Cutaneous exposure of adult Rana pipiens for 14 days to up to 20 mg/L atrazine resulted in a significant increase in ventilation rate, which may be indicative of respiratory distress. This effect was significant at the 12 mg/L atrazine concentration for thoracic ventilation and at 4.32 mg/L for buccal ventilation, although thoracic and buccal ventilation rates dropped at 7.2 and 12 mg/L atrazine, respectively (Allran and Karasov 2001). The estimated NOAEL for ventilation was 2.59 and 7.2 mg/L for buccal and thoracic ventilation, respectively. Frogs exposed to the highest atrazine concentrations (20 mg/L) showed no feeding response throughout the duration of the exposures but

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did not decrease in mass, possibly due to compensatory fluid gain from edema. The NOAEL for a reduction in feeding response was 12 mg/L (Allran and Karasov 2001).

8.2.2 Endocrine Disruption: Receptor Binding and Modes of Action Atrazine has demonstrated both estrogenic and anti-estrogenic activities in a number of in vivo and in vitro studies (Sanderson et al. 2001; Seung et al. 2003). Generally, there is little evidence that atrazine’s estrogenic or anti-estrogenic activities are mediated directly through the modulation of estrogen receptors. Through the use of in vivo exposure of immature female Sprague-Dawley rat uterii, and in vitro exposures of estrogen-responsive MCF-7 human breast cancer cell lines and the estrogen-dependent recombinant yeast strain PL3, Connor et al. (1996) demonstrated that the estrogenic and anti-estrogenic activities of atrazine were not mediated by the estrogen receptor. Neither atrazine nor simazine induced typical estrogenic responses in Sprague-Dawley rats in vivo (increased rat uterine wet weight, cytosolic PR binding, or uterine peroxidase activity), while both herbicides inhibited estrogen-induced PR binding and uterine peroxidase activity (Connor et al. 1996). Uterine estrogen receptor-binding capacity in ovariectomized rats fed triazines at 300 mg/ kg for 2 days was reduced by 30% (Tennant et al. 1994). However, atrazine was not able to competitively bind to receptors in the presence of estradiol, nor was there any displacement of ligand binding (Tennant et al. 1994; Roberge et al. 2004). The prevailing hypothesis is that at least some of the effects of atrazine are mediated through alterations in the expression or activity of cytochrome P450 19 (aromatase). Aromatase activity is modulated by phosphodiesterase (PDE), which converts cyclic adenosine monophosphate (cAMP) to 5’AMP; cAMP in turn can increase mRNA expression of aromatase (Sanderson et al. 2001). Atrazine can inhibit PDE, thus resulting in elevated concentrations of cAMP, which in turn increases the expression of aromatase (Sanderson et al. 2001; Roberge et al. 2004; Fan et al. 2007). Aromatase, in turn, converts testosterone to estrogen (and androstenedione to estrone). Aromatase has been induced by exposure to atrazine in vitro, in chelonian testis cell lines (Keller and McClellan-Green 2004), and human carcinoma cell lines (Sanderson et al. 2001). Nevertheless, there is still some potential for atrazine to demonstrate estrogenic or antiestrogenic activity independent of aromatase. First, estrogen receptors of nonmammalian taxa (i.e., reptiles or amphibians) may be more sensitive to atrazine. Vonier et al. (1996), for example, found that alligator (Alligator mississippiensis) estrogen receptors could bind with atrazine, whereas mammalian receptors typically do not bind with atrazine (Roberge et al. 2004). Both atrazine and cyanazine displaced [3H]17ß-estradiol from the α-estrogen receptor in alligators, and at 20.7 and 19 μM these 2 compounds inhibited 50% of the estradiol from binding to the receptor (Vonier et al. 1996). Thus, interactions between atrazine and the estrogen receptor are still possible, at least in nonmammalian taxa. Furthermore, Kniewald et al. (1995) found that atrazine inhibited 5α-dihydrotestosterone-specific receptor complex formation in rats exposed in vivo and in vitro. Similarly, the metabolite, deethyl­ atrazine, also inhibits 5α-reductase (Babic-Gojmerac et al. 1989) in a similar fashion as atrazine. Reduction of 5α-reductase in atrazine-exposed male rats decreases the conversion of testosterone to 5α-dihydrotestosterone (Kniewald et al. 1979), which is a potent androgen. Consequently, both atrazine and its metabolites can act as antiandrogens. Although atrazine did not affect estrogen binding to estrogen receptors, atrazine did reduce 5α-dihydrotestosterone binding to androgen-binding protein by about 40% (Danzo et al. 1997). Atrazine inhibition of 5α-dihydrotestosterone-specific receptor complex formation appears to be reversible, with the number of available binding sites for 5α-dihydrotestosterone returning to normal after 7 to 14 days postexposure in atrazine-treated rat prostates (Šimić et al. 1991). Dihydrotestosterone is not convertible to estrogen; thus, the inhibition of 5α-reductase by atrazine may increase the pool of testosterone available for conversion to estrogen. Exposure to exogenous testosterone, though not dihydrotestosterone, can cause feminization in turtles (Gutzke and Bull 1986), through the conversion to estrogen via aromatase.

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Furthermore, atrazine can inhibit 3 β-hydroxysteroid dehydrogenase in exposed rats (Kniewald et al. 1979), which catalyzes the conversions of pregnenolone, 17-hydroxypregnenolone, and dehydroepiandrosterone to progesterone, 7-hydroxyprogesterone, and androstenedione, respectively. The mechanism(s) of endocrine disruption by atrazine is not fully understood, but the current focus on atrazine’s interaction with aromatase should not be overemphasized to the exclusion of other viable hypotheses to its actions. Activity of 7-ethoxyresorufin (EROD) is directly associated with the induction of hepatic activity of the cytotchrome P450 1A1 enzyme. Similarly, O-demethylase (MROD) activity is another hepatic biomarker that is not as highly inducible as EROD but is more sensitive than some other monooxygenase enzymes in amphibians (Murphy et al. 2006c). Liver somatic index (wet weight of the liver divided by the total wet body weight of the fish multiplied by 100) can increase as the liver increases in size to allow greater detoxification of pollutants over long periods of exposure. In wild ranid frogs collected from agricultural and nonagricultural sites in Michigan, EROD and MROD activities were measurable in both adult and juvenile frogs and were similar among sites (Murphy et al. 2006c). Juvenile frogs had greater EROD and MROD activities than adult frogs. Rana catesbeiana and Rana pipiens had greater activities than Rana clamitans. Atrazine concentrations in water from the ponds were significantly and negatively correlated with MROD activity in adult male green frogs. Liver somatic index, EROD, and MROD activities of adult female and juvenile Rana clamitans were not significantly correlated with atrazine concentrations in water. However, liver somatic index values in adult male frogs differed significantly between agricultural and nonagricultural sites (Murphy et al. 2006c). 8.2.2.1 Adrenal Function The individual effects of cadmium (10 –8 to 10 –1 M), endosulfan, and atrazine (both at 10 –8 to 10 –4 M) on corticosterone secretion and viability of adrenal cells of Xenopus laevis and Rana catesbeiana were assessed using in vitro bioassay (Goulet and Hontela 2003). The ratio of the lethal concentration needed to kill 50% of the cells and the 50% effective concentration (LC50:EC50) was calculated with LC50 as the concentration that killed 50% of the steroidogenic cells and EC50 as the concentration that impaired corticosterone secretion by 50%. The higher the ratio, the greater the potential for endocrine disruption. Atrazine had no effect on cell viability and on corticosterone secretion in Xenopus laevis, but its endocrine-disrupting potential was high in Rana catesbeiana (Goulet and Hontela 2003). For comparison, LC50:EC50 ratios for cadmium and endosulfan in Xenopus were 26.07 and 1.23, respectively, and for atrazine, cadmium, and endosulfan in R. catesbeiana they were 909, 41, and 3, respectively, indicating that there was variation in species sensitivity, with atrazine being the most toxic chemical (Goulet and Hontela 2003). 8.2.2.2 Sexual Development The potential for atrazine to alter the sexual development, in particular gonadal development and associated sex steroid concentrations, has been studied in several species of amphibians, but primarily in Xenopus laevis. Studies have measured effects of atrazine concentrations in dose-response studies and field scenarios at environmentally relevant concentrations in the range of 0.1 μg/L to several hundred micrograms per liter. Commonly measured endpoints include gonadal anomalies such as testicular oocytes, intersex, hermaphroditism, decreased testosterone concentrations in plasma of male amphibians, alterations in gonadal somatic index, sex ratios, reduction in laryngeal muscle size, and aromatase activity. To examine the effects of atrazine on sexual development in amphibians, larvae of Xenopus laevis were exposed to a range of atrazine concentrations from 0.01 to 200 μg/L, or to untreated/ control conditions, throughout larval development. The cross-sectional diameter of larynges was significantly reduced in size in males produced from the exposed larvae at or above 1 μg/L atrazine (Hayes et al. 2002). All doses tested, except 0.01 μg/L atrazine, produced up to 20% occurrence of multiple gonads, whereas these abnormalities were not observed in the control animals. In sexually

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mature males, plasma testosterone concentratons were significantly lower and similar to testosterone concentratons in females when Xenopus laevis tadpoles were exposed every 3 days for 46 days to 25 μg/L atrazine (Hayes et al. 2002). The gonadal malformations induced by 0.1 to 25 μg/L atrazine in Xenopus laevis were then compared to those induced by an androgen receptor antagonist or 17 β estradiol. The combined frequency of multiple testes (single-sex polygonadalism), hermaphroditism, and nonpigmented ovaries was 10% or more in all concentrations of atrazine, whereas the frequency was less than ~2% in controls. Similar malformations were induced by estradiol except nonpigmented ovaries, which occurred only in the atrazine groups. The occurrence of gonadal malformations was significantly higher at all concentrations of atrazine compared to controls, but there was no clear linear dose-response at concentrations increasing from 0.1 to 25 μg/L (Hayes et al. 2006a). In 2003, Carr et al. reported no effects on sex ratio in larval Xenopus laevis exposed to 0, 1, 10, or 25 μg/L atrazine at 48 hours posthatch until metamorphosis. At 25 μg/L atrazine, the incidence of intersex animals increased to 4.7%, which was significantly different than controls, and the incidence of intersex increased with atrazine concentrations. Atrazine at the tested concentrations did not reduce the size of the laryngeal dilator muscle, which is sexually dimorphic in Xenopus laevis. In contrast, Xenopus laevis tadpoles were also exposed under static conditions to atrazine (21 μg/L) for 48 hours during sexual differentiation (Tavera- Mendoza et al. 2002). The histology of the gonads indicated 57% reduction in testicular volume among atrazine-exposed tadpoles, primary spermatagonial cell nests were reduced by 70%, and nursing cells that provide nutritive support for the developing germ cells declined by 74%. Testicular resorption was observed among 70% and failure of full development of the testis occurred in 10% of atrazine-exposed tadpoles. In a longer-term study, Xenopus laevis was exposed to 0.1 to 25 μg/L atrazine from 72 hours posthatch until 2 to 3 months postmetamorphosis in a 3-day static renewal system (Coady et al. 2005). Mortality, growth, gonadal development, laryngeal muscle size, and aromatase activity in juveniles were not affected. Male frogs exposed to 1 μg/L atrazine had lower estradiol concentratons, although there was not a consistent dose-response. Similarly, male Xenopus laevis exposed to 10 or 100 μg/L atrazine for 49 days showed neither gonadal abnormalities nor significant differences in plasma estradiol or testosterone concentrations. The gonadal somatic index was significantly higher in males exposed to 10 μg/L atrazine, but there was no consistent dose-response (Hecker et al. 2005a). Also, male Xenopus laevis were exposed to 1, 25, and 250 μg/L atrazine for 36 days and testicular aromatase activity and cytochrome P450 19 (aromatase) gene expression, testosterone, and 17ß-estradiol and gonad size were measured (Hecker et al. 2005b). Aromatase gene expression was measured because at relatively high concentrations atrazine can upregulate P450 19 gene expression in the human adrenocarcinoma cell line (Sanderson et al. 2001). There were no effects found at any concentration except testosterone concentrations in plasma, which were significantly lower in males exposed to 250 μg/L atrazine (Hecker et al. 2005b). Effects of atrazine on sex differentiation were studied using “wild-type” Xenopus laevis tadpoles and all-ZZ male cohorts of X. laevis, produced by mating wild-type ZZ male to sex-reversed ZZ male (female phenotype) (Oka et al. 2008). Stage 49 (Nieuwkoop and Faber 1994) tadpoles were exposed to 0.1 to 100 μg/L atrazine or 0.27 μg/L 17ß-estaradiol induction until all larvae completed metamorphosis. Atrazine had no effect on metamorphosis of developing wild-type or all-male X. laevis larvae. A statistical increase in female ratios was observed in 10 and 100 μg/L atrazine exposures relative to a control group. However, no hermaphroditism or sex reversal was found, and P450 aromatase mRNA in the gonad and hepatic vitellogenin were not induced in the atrazine treatments. The authors suggest that effects of atrazine on sexual differentiation were not caused by estrogenic action, and hasatrazine had no induction ability of P450 aromatase in the gonad (Oka et al. 2008). To assess the transgenerational effects of atrazine, reproductive success and development of F2 offspring from F1 adult Xenopus laevis exposed to atrazine throughout larval development and as sexually mature adults were tested using larvae exposed to 1 of 4 nominal concentrations of

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atrazine (0, 1, 10, and 25 μg/L) beginning 96 hours after fertilization and continuing through 2 years postmetamorphosis (Du Preez et al. 2008). Clutch size and survival of offspring were used as measurement endpoints of reproductive success of the F1 frogs. Larval survivorship and time to metamorphosis were used to measure developmental success of the F2 offspring from atrazine-exposed frogs. Testes in F1 and F2 frogs were examined for incidence of anomalies, such as testicular ovarian follicles, and sex ratios in F2 offspring were investigated to determine if exposure to atrazine caused transgenerational effects (effects on F2 individuals due to exposure of F1 individuals). There were no effects from any of the concentrations of atrazine on clutch size of F1 frogs. There were also no effects on hatching success or time to metamorphosis. Sex ratios did not differ between F2 offspring among treatments. There was no evidence to suggest a transgenerational effect of atrazine on spawning success or reproductive development of Xenopus laevis. However, in atrazine-treated male Rana pipiens (0.1 and 25 μg/L) rates of 36% gonadal dysgenesis (underdeveloped testes with poor structure and closed lobules and low to absent germ cells) were found at 0.1 μg/L, and the rate was 12% at 25 μg/L. At 0.1 μg/L, 29% of the animals exhibited varying degrees of sex reversal, while only 8% showed this response at 25 μg/L. Sex-reversed males contained oocytes in testes while control males showed no oocytes in testes; 2 control males showed degenerating extragonadal oocytes and 1 had gonadal dysgenesis (Hayes et al. 2003). Rana pipiens were exposed to 10 mg/L nitrate or 10 μg/L atrazine or combined exposure of both of these treatments and compared to controls (Orton et al. 2006). Testicular oocytes were found in nitrate-only, atrazine-only, and the control groups, but not in the combined treatment. In the atrazine-alone, nitrate-alone, and combined treatments, there were a decreased percentage of spermatocytes, an increased percentage of spermatids in testes, and an increased mature ovarian follicle size in females (Orton et al. 2006). In a dose-response study conducted using a protocol developed with and approved by the USEPA (Hosmer et al. 2007; Steeger et al. 2007), exposures of Xenopus laevis to 0.01 to 100 μg/L atrazine were conducted under flow-through conditions and no testicular oocytes or intersex cases occurred in atrazine exposed groups (Hosmer et al. 2007; Steeger et al. 2007). The use of flow-through conditions contrasts with methods used in most previous studies of atrazine exposure and effects on gonadal development in amphibians, in which static renewal conditions were used, making comparisons among results of studies difficult. Flow-through conditions are generally not representative of conditions in small farm ponds and ditches, where amphibians are most likely to experience the exposure without substantial water flow. A number of field surveys of frogs collected in agricultural areas (Reeder et al. 1998; Hayes et al. 2003; Hecker et al. 2004; Jooste et al. 2005), especially those areas where corn production and atrazine use were dominant, have found physical and physiological abnormalities that are consistent with those reported in laboratory dosing studies with atrazine, although many of these surveys have failed to find a dose-response relationship specific to environmental atrazine concentrations. While field studies are limited in the sense that they cannot be used to establish a cause-effect relationship, they are useful in exploring whether the effects seen in the laboratory are apparent in or relevant to wild populations. Several investigations have examined the occurrence of atrazine effects in the native, wild populations of Xenopus in South Africa. Xenopus laevis were collected from wild populations in corn-growing regions with relatively typical surface water exposures of atrazine (greatest mean concentrations among sites of 3.5 to 4.1 μg/L) and non-corn-growing regions (Hecker et al. 2004). The trends in the combined data among sites showed negative correlations between plasma testosterone and atrazine exposure and its degradation products deisopropylatrazine, deethylatrazine, and tertbuthylazine in females, and between testosterone and diaminochloroatrazine in males. Estradiol in females exhibited a significant negative correlation with atrazine and deethylatrazine. No correlations were found between gonadal aromatase activity or gonadal somatic index and any chemicals measured. In the same study areas, there were no significant differences in laryngeal mass in X. laevis males from corn-growing and non-corn-growing areas (Hecker et al. 2004). Mean percent

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fractional volume of seminiferous tubule distribution of testicular cell types, such as percentage of spermatogonia, spermatocytes, and spermatozoa, was not different among sites, nor was the incidence of testicular oocytes (Smith et al. 2005). Xenopus laevis larvae were exposed to atrazine concentrations of 1, 10, and 25 μg/L in microcosms until just after metamorphosis (Jooste et al. 2005). They exhibited no significant increase in incidence of testicular oocytes relative to controls, although the rate in the control group was 57%. The mean number of testicular oocytes per individual was 9.5 in the control group and 11.1 in the 25 μg/L treatment group. Ten months after metamorphosis, another subset of juveniles was examined and the maximum number of testicular oocytes found was 5 per individual (Jooste et al. 2005). In the early 1990s, species native to North America were starting to be examined in the wild in the context of pesticide exposure and effects on sexual development. Cricket frogs (Acris crepitans) from 2 general areas (study A and study B), including several sites each, in Illinois were collected to assess the effects of environmental contamination on the prevalence of intersex gonads and sex ratios. Of 341 frogs collected (1993 to 1995), 2.7% were intersex individuals (Reeder et al. 1998). There was no significant correlation between chemical compounds detected (65 herbicides, insecticides, and fungicides; 13 organochlorine pesticides; total polychlorinated biphenyl (PCBs) and polychlorinated dibenzo-furans (PCDFs), and dioxins; and lead and mercury in sediments and water) and cricket frog intersexuality. However, in study A, in 1 year (1994), there was a weak association approaching significance (p = 0.07) between the detection of atrazine and occurrence of intersex individuals. Pesticides detected in sediment included atrazine, cyanazine, and metolachlor (where detected concentrations ranged from 3 to 17 μg/L), and chlorpyrifos was detected in water at 3.1 mg/L. In study B, the sex ratios of juvenile frogs differed significantly between PCB/PCDF/ PCDD-contaminated sites and reference sites (n = 16 frogs collected per site). In a 2001 field study of Rana pipiens collected from ponds in 8 agricultural sites across a gradient of atrazine concentrations up to 12 μg/L (atrazine application rates in these areas ranged from 0.4 kg/km2 to >28.7 kg/km2), all sites associated with atrazine rates of more than 0.4 kg/km2 and/ or higher than 0.2 μg/L in water contained male Rana pipiens that displayed sex reversal similar to that found in dose-response studies on the same species (Hayes et al. 2003). However, there was no linear dose-response association of increasing gonadal malformation and concentrations of atrazine and its breakdown products at the time of sampling (i.e., sites with the highest atrazine concentration did not exhibit the highest rates of abnormalities in gonads among the 8 study areas). There were other pesticides used in these sites and analysis was conducted on water for these chemicals. However, atrazine and its metabolites were detected at all sites, but among the other pesticides, only metalochlor was detected, and only at 1 site (Hayes et al. 2003). During the field collections juvenile Rana pipiens were abundant at all of the collection sites, suggesting that the gonadal anomalies may be reversible, some percentage of the population is unaffected, these morphological anomalies do not impair reproduction, or resistance to atrazine may occur in the population (Hayes et al. 2003). Genetic variation within populations of amphibians can explain a significant amount of the variation in tolerance to insecticide (e.g., carbaryl) (Semlitsch et al. 2008), thereby suggesting it is a heritable genetic trait that could contribute to resistant populations of amphibians arising in agricultural areas. In 2002 and 2003, frogs were collected from field sites in Michigan where atrazine occurred (nondetectable to 2 μg/L at all but 1 site, which had a concentration of 250 μg/L measured in 1 year of the study). In these field exposures, aromatase activity was measurable in less than 11% of testes in adult male Rana clamitans, and less than 4% of testes in juvenile males. Atrazine concentrations in field sites did not correlate with aromatase activity or with plasma steroid concentrations in Rana clamitans (Murphy et al. 2006b). In the family Bufonidae, male toads possess rudimentary ovaries, called Bidder’s organs, which are attached to the testes (Pancak-Roessler and Norris 1991). In field collections of Bufo marinus from sugarcane- and non-sugarcane-growing areas and urban sites in Florida, toads were examined for the presence of developed Bidder’s organs in 3 separate studies, which found similar trends. In

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the first study (Sepulveda and Gross 2003), there was an increased incidence of intersex (ovarian tissue in Bidder’s organ) in toads identified as having testes. Approximately 29% of males from one agricultural site (Belle Glade) and 39% of males from another (Canal Point) were intersexed, while no intersex frogs were identified among the toads from a nonagricultural site. Vitellogenin, a female-specific protein, was present in intersex toads at concentratons similar to those in females and was about double that in male toads (Sepulveda and Gross 2003). Atrazine concentrations in the agricultural site range from